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 Water2 Table of Contents:

Page Executive Summary

--------------------------------------------------------------------- 3

Introduction

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

------------------------------------------------------------ 5 Effects on Fish Population

------------------------------------------------------------- 7 Physiologi cal 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

Con clusion

------------------------------------------------------------------------------ 30

References

------------------------------------------------------------------------------- 313 Executive Summary :

The issue of Suspended and B edded 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 wa ter 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 artifi cial (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 aquati c 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 Soc iety to be of special concern, threatened, or endangered . Among North American freshwa ter species, approximately 37% of fishes, 35% of amphibians, 73% of mussels, and 65% of crayfish have been classified as vulnerable, imperiled, or endangered. This re view 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.

Ph ysiological 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 a n adverse effect on fish health, metabolic demand, and g rowth 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 eff

ect 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

part icles, 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, a nd reduced larval survival These impacts are associated with amount of elevated sediment loads, size4

of the sediment particles involved, seasonal variation, rates of sediment deposition and reduction in oxygen supply.

There are multiple ecological effec ts 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 betwee n 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 . E ven 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 rat es, 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

re sult 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 ch annel 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 fish5

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 du st 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 bo dies 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 syn thesized 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 cons idered 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 ca tegorized 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 ide ntified 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 parti cle 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 impa irment 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 ri vers 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 including6

sedimentation. Among North American freshwa ter specie s, 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 d rastic declines have occurred i n 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 endan gered 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 problem s 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 remova l 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

environment s 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

capa bilities 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 i mbalance 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 docume nted 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 severit y 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 spec ies 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 water

filled 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 and7

fau na. 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 plant s. 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 du ration 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, short

term 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 re port, 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

dos es 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 an d 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 an d 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, although8

responses to the three lowest concentrations were not significantly differ ent. 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. Ho wever, 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 inc reased 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 traumati zation 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, erythr ocyte 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. Newcom be (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 27 0 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 fo r 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 j uvenile 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 stressor s 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

repo sitory 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 infestation s 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 che mical 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

suspe nded 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 increas ed 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, s ome 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 condition s, 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 turbid ity 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 (19 77) 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 (Onc orhynchus 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 turbi d 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 red uce 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 long

lasting 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 of10

disturbance, the level of stress (and associated ene rgy 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 (199 3) 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 con centration 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 en ergy 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 habit at 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, th e 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 su spended 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, tempera ture, 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 mort ality 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 po int 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. Howeve r, 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 e mphasis 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. Th ey 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 estima tion 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 biologi cal 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 habitat s, 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 int erruption 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 s tream 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 redu ce 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

model s 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 th e 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 sapidi ssima ), 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 salmonid s 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 equation s, 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 inse cts 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 h as13

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. (19 77) 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 o r 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 clear

cut logging pract ices. 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

cleari ng 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 r egions. 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 in dicated 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, unsilted14

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 coat ed 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 i nferred 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 c ause 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/cm 2 (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. Thi s 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 (199 9), 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 whe n 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) sugges ted 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 t hat hatching success of coho salmon and cutthroat trout15

(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 o r 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 potentia l 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 inter stitial 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 i n 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 graduall y 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

intrus ion 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 early16

e mbryonic 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 macro

invertebrates require the interstitial spaces at the bottom of streams to lay their eggs. Most fast flowing, lower order streams l ike 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 interstic es 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 sedime nt 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 th e 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 logg ing 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

literat ure 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 inverteb rates 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 (19 61) 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 fil tration 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 compositi on 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 s hift 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 systematica lly 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 benthi c 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 inverteb rates. 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

- inverteb rate 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 influenc es 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 1 5 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 (Ephemeropte ra, 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 field18

experiment reduced some confounding e nvironmental 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 su rvey 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 a bundances 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

str eams 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 o f 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 Brus ven (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

ma ting 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 res ources 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 diversit y 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)

hav e 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 st rong 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 o f species, and ratio of EPT to chironomids, all indicators of a pollution19

intolerant assemblage, are negatively associated with embeddedness and positively

associated with width

- to

- depth ratio and percent riffle (Nerbonne and Vondacek 2001). Zweig and Rabe ni, (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 r epresenting 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 m acroinvertebrate density and diversity are directly related to substrate diversity

. Lenat et al (1981) reported that as sediment settles, inters titial spaces between coarse substrata are filled, which reduces available habitat for macro

invertebrates. If enough sediment fills interstitial spaces, then an impermeable sediment barrier may form that diminishes available living habitat and causes redu ctions 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% reductio n 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). Wagene r 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) detec ted 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 r eported 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

comm unity 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/L20

resulted in a 25 to 60% decrease in the density of stream macroinvertebrates. Likewise, Slaney et al. (1977) found that a 16 hour p ulse 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 Chattoo ga 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 ecol ogy 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 suspended

sediment 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 macro

invertebrates in the Blue Ridge physiography. Corre sponding 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 managem ent 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 e cosystems. 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 th at 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 inve rtebrate 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 an21

altered habitat. So uthwood (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 successiona l 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 recover y, 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 t o 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 bedde d 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 ma trix 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 aq uatic 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 p articles 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 mussel s 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) i n an attempt to d etermine 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 si lt (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 speci es.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 sele ction efficiency for organic matter. Newell and Jordan ( 1983) demonstrated that t he 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 biv alves. 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 e fficiencies 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 inhabit ing 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 s ediment, 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 exa mple, 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 cra yfish 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 availabilit y 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 macrochi rus), 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 m ouths. 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 most24

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 sedime nt 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 ene rgy 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) r effered some c lassic 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 an d 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 c ommercially 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 on e 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 - 2 d - 1 . Reduced growth

ra tes 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 - 2 d - 1 (r 2 =0.92). Their study indicated that

s tress 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 symbioti c 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 aquatic

dependent wildlife (e.g., birds and seals). Reede r (1951) and Fritzell et al. (1979) suggested that wildlife habitat is the important component of wetland and a n important25

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 item s 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 mor e 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 wide

spread and long

- term, however, they might cause a problem for aquatic

- dependent wildlife that consume aquatic life. For example Mc Intyre (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 rel ationship 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 in crease 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 t hat 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 correlative

type studies, e specially 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 ne eded 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 sy stems 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) indic ated 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 sub merged 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 d ecreased 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 se diment 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 redu ced 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 sedim ent 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.

Stei nman 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 penetr ation 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 redu ction 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. A quatic 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 of27

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 aqua tic 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 f

or 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 pr oductivity 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 addr ess 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

tha t 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 a gricultural, 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

excess ive 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

cause s 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 incre ased sedimentation on lotic food webs is pervasive and begins at primary trophic levels. Increased rates of sedimentation and28

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. (1 994) 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 pr airie 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 germ inate 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 div ersity 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 pla nts, 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 field s. 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) in dicated 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 soil

water 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 com bined with differences in light intensity, high rates of sediment accumulation had slight negative effects under natural29

light, but under shaded conditions these negative effects were substantially increased, causing changes to the structure of the whole a ssemblage. 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, an d 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 a re 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, c hannel 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 Sc orgie (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 disturbanc e 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 m otorboat 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, t urbulence, 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 compositio n 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 a30

water bodies exceeds the capacity of the watercourse to transport or assimilate the

sediment str ess 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

r esource 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 o r 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 th e 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 dec reased 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) o bserved 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 nu merous 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 bodi es 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 natura l 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 aquat ic 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 presente d 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 expres sed through pervasive alterations in local food31

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 av ailability 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 ra tes of growth, reproduction, and recruitment. I mbalance 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

whic h water quality criteria for sedimentation have yet to be established.

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