United States EPA-600/3-81-013
Environmental Protection February 1981
Agency
v>EPA Research and
Development
Sediment Removal as
a Lake Restoration
Technique
Prepared for
Office of Water Regulations and
Standards
Criteria and Standards Division
Prepared by
Environmental Research Laboratory
Office of Research and Development
Corvallis, OR 97330
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EPA-600/3-81-013
February 1981
SEDIMENT REMOVAL AS A LAKE RESTORATION TECHNIQUE
by
Spencer A. Peterson
Freshwater Division
Corvallis Environmental Research Laboratory
Corvallis, Oregon 97330
CORVALLIS ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
CORVALLIS, OREGON 97330
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DISCLAIMER
This report has been reviewed by the Corvallis Environmental Research
Laboratory, U.S. Environmental Protection Agency, and approved for publica-
tion. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
ii
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ABSTRACT
Sediment removal as a lake restoration technique is reviewed to examine
its positive and negative aspects. The effectiveness and longevity of the
process is considered in terms of retarding eutrophication and reducing the
impacts of toxic sediments. Freshwater lake sediment removal is usually under-
taken to deepen a lake thereby increasing its volume to enhance fish produc-
tion, to remove nutrient rich sediment, to remove toxic or hazardous material,
or to reduce the abundance of rooted aquatic plants. Review of more than 60
projects and examination of five case histories reveals that the first three
objectives are usually met through sediment removal. The effectiveness of
dredging to control aquatic plants has not been well documented.
Advantages of sediment removal techniques include ability to selectively
deepen part of a lake basin, increase the lake volume, recover organically
rich sediment for soil enrichment, and improve limnetic lake quality. Dis-
advantages include high cost, phosphorus release from sediment, increased
phytoplankton productivity, noise, lake drawdown, temporary reduction in
benthic fish food organisms, the potential for release of toxic materials to
the overlying water and for environmental degradation at the dredged material
disposal site. High quality dredge material can be used for beneficial pur-
poses and may offset the initial high cost of dredging.
The technique is recommended for deepening and for reducing phosphorus
release from sediment. Sediment removal to control toxic materials is pos-
sible with minimal environmental impact when proper equipment is used, but it
may be extremely expensive. Dredging will remove rooted aquatic plants,
however, their re-encroachment rate will be depth, sediment texture, and sedi-
ment nutrient dependent. Lack of definitive information about the effect of
these factors on the regrowth of various plant communities prohibits explicit
recommendations on sediment removal to control rooted plants.
ill
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CONTENTS
I. Introduction 1
II. Purposes of Sediment Removal 1
III. Considerations for Sediment Removal 5
IV. Case Histories of Sediment Removal 20
V. Costs of Sediment Removal 29
VI. Summary 29
VII. References 32
VIII. Appendix: Status of Lake Sediment Removal Projects 42
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ACKNOWLEDGEMENTS
I thank Joanne Oshiro of the Corvallis Environmental Research Laboratory,
(CERL) USEPA, Corvallis, Oregon for her initiative and persistence in obtain-
ing and assembling the data in the Appendix table of this paper. I also thank
those who assisted her in this task by supplying unpublished information.
They include Regional Environmental Protection Agency Clean Lakes coordina-
tors, State personnel and project managers too numerous to mention. Thanks
goes also to Phil Larsen of CERL who offered constructive criticism of the
manuscript and suggestions for its improvement.
vi
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I. Introduction
While control of external pollutant inputs to a lake should be given high
priority in any restoration/protection plan, external controls alone may be
insufficient to produce the desired changes in lake quality (Emery et aj. ,
1973; Minder, 1948 in Peterson et aj. , 1976: Larsen et aj., 1975). Where lake
improvement goals cannot be met by watershed nonpoint source pollutant con-
trols alone, sediment removal might be necessary.
Sediment removal usually includes one or more of four basic objectives.
The first is simply to deepen a lake to improve boating, fishing, water skiing
and other uses impaired by the effects of shoaling. A second objective is to
prevent or reduce the internal cycling of nutrients. It has been shown that
sediment recycling of nutrients in lakes may represent a significant fraction
of the total nutrient loading to the lake (Funk, 1978; Larsen ejt al. , 1975;
Larsen et aj. , 1979; Cooke e_t afL , 1977). A third objective might be to
remove toxic sediment. A fourth objective, or possibly a fringe benefit of
the first and second objectives, is to reduce nuisance aquatic macrophyte
growths.
Until recently there was little documentation of the effectiveness of
sediment removal for meeting these goals. Pierce (1970) reviewed 49 lake
sediment removal projects and concluded that none provided sufficient data to
determine the effects of lake dredging on the total lake environment. Four
years later, Dunst et aj. (1974) reported on 50 world-wide lake sediment
removal projects, again pointing out a general lack of documented effects.
Since 1974, additional lake sediment removal projects have been initiated.
Some are complete with documented results while others are ongoing. The
purpose of this paper is to review and update information on the effectiveness
of sediment removal as a lake restoration technique. Personal communications
are referenced liberally where documented results are unavailable and lakes of
the United States are emphasized. The effectiveness of different types of
sediment removal techniques and various objectives are illustrated by five
case histories.
II. Purposes of Sediment Removal
1. Deepening
Sediment removal for maintaining navigation channels has been practiced
for years. These deepening projects generally have been successful (Herbich,
1975). However, most have failed to address the source of the problem in the
watershed, thus they necessitate rather frequent routine maintenance work. In
small lakes, sediment removal may become necessary when uses such as boating
are impaired as a direct result of shoaling. Other considerations pertaining
to deepening are that lake volume must be sufficient to offset loss by seepage
and evaporation. The U.S. Department of Agriculture (1971) recommends small
lake and pond depths ranging from 1.5 m to 4.5 m for various parts of the
country to compensate for water loss via these mechanisms. Toubier and
Westmacott (1976) indicated that lakes in colder parts of the United States
must be at least 4.5 m deep to avoid winter fish kills. Each of these con-
siderations addresses sediment removal strictly from the standpoint of need to
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deepen the lake basin. Pierce1s (1970) survey of upper Midwestern dredging
projects showed that projects undertaken for the purpose of deepening were all
successful. The conclusion is not surprising, for who can argue that if
sediment is removed from a lake basin, that the lake will be deeper?
2. Nutrient Removal
Nuisance algal growth and nutrient relationships in lakes are well docu-
mented; phosphorus is most frequently cited as the limiting nutrient (Bartsch,
1972; Porcella et a_L , 1974; Schindler, 1977). Uptake of nutrients from the
water and sediment by plants ultimately results in deposition of organic
materials into the lake basin. Table 1 shows the total phosphorus content of
sediments in selected lakes in North America. A portion of this phosphorus
may be released during spring and fall lake circulation in dimictic systems.
In shallow polymictic lakes, sedimentary phosphorus release may be more fre-
quent, constituting a greater nuisance by periodic infusion of nutrients to
overlying water during summer recreation periods. The periodic influx of
nutrients in this manner usually results in an over-abundant growth of
unwanted phytoplankton causing reduced water transparency, undesirable green
water color, and in severe cases a serious depletion of dissolved oxygen and
fish kills. Livingston and Boykin (1962) estimated that sediment-regenerated
phosphorus amounted to approximately 45% of the phosphorus loading to Linsley
Pond, Connecticut. Welch et a]_. (1979) estimated the phosphorus loading from
sediments in Long Lake, Washington to be 200-400 kg yr-1 or about 25%-50% of
the external loading. Jones and Bowser (1978), citing a personal communica-
tion from Callender (1978), reported that sediment phosphorus regeneration in
White Lake, Michigan constituted approximately 40% of that lake's phosphorus
loading. Larsen and Schults (1978) stated that, "the supply of phosphorus
from sources within a lake can be several times the supplies from external
sources for periods of time on the order of months." Shagawa Lake, Minnesota
experiences internally recycled summer phosphorus pulses of approximately
2000-3000 kg during June, July and August. This compares to an annual loading
of phosphorus to the lake from wastewater (before advanced waste treatment) of
5000-5500 kg, and current loading of about 1000-1500 kg (Larsen, 1980). Thus,
internal phosphorus loading to Shagawa Lake prior to advanced waste treatment
(AWT) was about 28-35% of the total loading. Following implementation of AWT,
internal phosphorus loading constituted approximately 66% of the total load-
ing. Continued internal nutrient loading to the lake has been responsible for
the apparent slow recovery rate.
It is evident that internal phosphorus loading to a lake can, in some
instances, amount to major portions of the total loading. Thus, sediment
removal to prevent internal loading should be weighed against alternative
treatment techniques and given serious consideration where it might be
expected to produce significant loading reductions.
3. Toxic Substances Removal
Sediment contaminated with toxic and hazardous materials poses a poten-
tially severe problem for highly industrialized countries throughout the
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Table 1. Total phosphorus content of sediments
America (from Larsen and Schults, 1978)
in selected lakes in North
Lake
Total Phosphorus
(mg/g dry wt)
Core Depth
Reference
Sammamish, WA
Lower St. Regis, NY
Huron
Great
Ontario Lakes
Erie
2 - 5
0.5 - 1.4
M.2 - *3.i
0.19 - 2.9
Monona, WI
Washington, WA
Shagawa, MN
- >6
1 - 5
40 cm core
35 cm core
50 cm core
50 cm core
50 cm core
top 3 cm
from 48 sites
100 cm core
40 cm core
160 cm core
Welch (1977)
Fuhs et al_. (1977)
Kemp et al_. (1972)
Kemp et al_. (1972)
Kemp et a].. (1972)
Williams et al.
(1976)
Bortleson and Lee
(1975)
Shapiro et al.
(1971)
Bradbury and
Waddington (1973)
world -- "potential" because in many cases these problems simply have not been
recognized. This picture is changing, however, as more thorough and sophisti-
cated lake sediment surveys are conducted (Matsubara, 1979; Sakakibara and
Hayashi, 1979; Horn and Hetl ing, 1978; Bremer, 1979; Mackenthun et al_. , 1979).
Theoretically, toxic materials in sediment might be controlled by cover-
ing or in-place detoxification. These treatment procedures are relatively
untested. The obvious alternative is to remove the polluted sediment. The
question is how to remove it without reintroducing toxicants into the water
column and without causing secondary pollution problems at the disposal sites.
Turner and Fairweather (1974) stated that "Once a body of water is polluted,
the most effective means known to man to remove bottom sediment pollutants is
an efficient, hydraulic dredge". Emphasis here should be placed on the word
"efficient". While hydraulic dredges are the workhorse of the dredging indus-
try, and are effective in moving large volumes of dredged material with mini-
mal energy input, they are not always effective in minimizing the resuspension
of bottom sediment. The key to removing polluted sediments is in minimizing
sediment-water interface disruption, containment of polluted dredge materials
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and treatment of decant water. Several mechanisms for attaining these goals
have been described (Peterson and Randolph, 1977, 1978, 1979; Peterson, 1979;
Barnard, 1978) and will be discussed later in this paper under sediment.
removal methods.
4. Rooted Macrophyte Removal
Rooted aquatic plants in the littoral zone of lakes often interfere
directly with recreational uses. Fishing, boating and swimming become diffi-
cult or nearly impossible; further, an over-abundance of rooted plants may be
aesthetically displeasing to some persons. Any one of these problems may be
sufficient justification for removing part of the aquatic plant beds. Dredg-
ing is but one of several approaches, however, due to cost and convenience
differences, harvesting and chemical treatments are more commonly used to
"solve" these direct use impairment problems (Robson, 1974).
While direct use impairment problems may justify the removal of some
macrophytes from lakes, there is an increasing literature on the effects of
macrophytes on internal nutrient cycling. Their role in this process, with
its effect on algal dynamics, may be an even more important reason for attempt-
ing to control macrophytes by removing them from a lake. Wetzel (1975) indi-
cates that most of the organic matter found in small lakes may be derived from
their littoral zone. Reimold (1972) showed that salt marsh plants (Spartina)
act as nutrient pumps, extracting sediment nutrients, translocating them to
stems and leaves and excreting them to the surrounding water on each succes-
sive tidal cycle. Other researchers have documented that various species of
freshwater aquatic plants extract nutrients chiefly from the sediment and
translocate them to the surrounding water (Schults and Malueg, 1971; Twilley
et aj_. , 1977; Carignan and Kalff, 1980).
There is now considerable evidence that healthy freshwater aquatic macro-
phytes do not excrete large quantities of nutrients to the surrounding water
while in the active growth phase. They do, however, tend to concentrate
sediment supplied nutrients in their tissues. Some loss is incurred through
plant part sloughing, but major amounts of nutrients are recycled to the lake
when plants fruit and during senescence and plant decay (Lie, 1979; Welch et
a]_. , 1979; Barko and Smart, 1980). Barko and Smart (1979) estimated that
in-lake mobilization of phosphorus by Myriophyllum in Lake Wingra, Wisconsin
might amount to approximately 60% of the annual external phosphorus loading.
Welch et aj_. (1979) indicated that much of the "sediment" phosphorus loading
in Long Lake, Washington was probably due to rapid plant die-off and decay.
While macrophyte removal may be incidental to the objectives of many lake
deepening and nutrient control projects, it appears likely that the removal of
these plants may be far more important to reducing in-lake nutrient cycling
than had been suspected previously. Current evidence indicates that any long
range lake restoration project concerned with in-lake nutrient controls will
need to focus on both the macrophyte and sediment compartments (Carignan and
Kalff, 1980; Barko, 1980).
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III. Considerations for Sediment Removal
Currently there is no single, off the shelf, commonly accepted problem
assessment procedure, remedial action plan, preferred piece of equipment or
even agreement about when, for what reason, or how to remove sediment from a
freshwater lake. Nevertheless, there are a number of factors which must be
weighed when considering lake remedial action. Some are more critical to
plans for sediment removal than they might be for other types of action.
Factors which must be considered for any sediment removal project are: 1)
determination of problem sources; 2) characterization of sediment; 3) determi-
nation of sediment removal depth; 4) environmental problems associated with
sediment removal; 5) sediment removal methods; 6) sediment disposal area; and
7) lake conditions most suitable for sediment removal.
Clear-cut, definitive statements about each of the above factors cannot
always be made. In some cases this may be due to basic lack of information.
In others it will depend on individual lake circumstances such as the intended
use of the lake, the geographic and socioeconomic setting of the area, and the
project completion time-frame. The following discussion addresses each of the
above factors.
1. Determination of Problem Sources
Lake problems generally are identified as a result of some use impair-
ment. The use impairment problem may be identified by nothing more than the
casual observation that the lake becomes excessively green with algae or
choked with rooted aquatic plants at various times of the year. In the case
of over-abundent algal growth the problem is most frequently traced to exces-
sive amounts of phosphorus. Chapra and Tarapchak (1976) presented a cause and
effect sequence of events related to phytoplankton problems in eutrophic
lakes. The approximate sequence is that increased phosphorus loading produces
increased mean annual phosphorus concentrations in water which lead to
increased mean summer concentrations of chlorophyll a, which lead to increased
water turbidity, increased Secchi depths, and finally to reduced dissolved
oxygen levels when sedimented organic material decomposes. Macrophytes pre-
sent a similar problem with regard to dissolved oxygen depletion and potential
for fish kills. They differ, however, with respect to nutrient supply sources
since most are derived from the sediment (Carignan and Kalff, 1980; Barko and
Smart, 1980).
While casual observations of lake conditions may serve as the catalyst
for action, they are insufficient for the development of lake restoration
plans. More quantitative data are required to determine causes of perceived
problems. One of the most common and useful approaches is a determination of
nutrient budgets because algal and associated problems have been shown
repeatedly to correlate with the nutrient level of lakes (Vollenweider, 1968;
Vollenweider, 1976; Dillon, 1975; Larsen and Mercier, 1976; Dillon and Rigler,
1974). Reckhow (1979) critically reviewed a number of nutrient mass balance
approaches and concluded that several will perform satisfactorily in predict-
ing trophic state for a given set of lakes. However, Vollenweider1s (1975)
loading criteria have been most generally used due to ease of application in
assessing the susceptibility of a lake to altered phosphorus loading.
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Mass balances focus on methods to account for surface and subsurface
nutrient inputs and outputs. They do not account directly for internal nutri-
ent recycling, but can be used to estimate the relative importance of internal
nutrient supplies. The internal gain or loss of nutrients can be calculated
as the difference between the algebraic sum of the inputs and outputs of the
lake and the change in nutrient concentration in the lake over the same time
frame. Figure 1 illustrates the potential importance of internal nutrient
cycling in shallow lakes and indicates how nutrient washout model predictions
of lake recovery might vary from measured in-lake results.
OBSERVED
HYDRAULIC WASHOUT MODEL
PHOSPHORUS WASHOUT MODEL
1973
1974
1975
Figure 1. Comparison of hydraulic and phosphorus washout models with observa-
tions in Shagawa Lake, 1973-1974. Total phosphorus concentrations
are the averages for the entire lake (updated from Larsen et al. ,
1975).
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Larsen et a\_. (1979) used a modified form of the phosphorus residence time
model to account for and accurately describe periodic summer pulses of inter-
nal phosphorus loading in Shagawa Lake, Minnesota. Application of modeling
techniques or direct measures of nutrient release rates from sediment and
calculation of the sediments' nutrient contribution to the lake is an impor-
tant step toward determining the relative importance of the sediment compart-
ment to the overall loading of the lake. Scheider et a]I. (1979) have summar-
ized methods for using hydraulic and nutrient budgets to assess lake problems.
A more recent and comprehensive review of the subject is presented by Reckow
(1980).
2. Characterization of Sediment
If sediment removal appears to be a reasonable lake restoration option
after an initial analysis, it may become necessary to further refine the
sediment information base. It would be useful to determine various physical
characteristics such as sediment depth and distribution, particle size and
organic content, bulk density, and other factors critical to sediment removal.
Chemically one might want to know if the sediment contains toxicants, what the
nutrient content is and what the rate of release, of nutrients from the sedi-
ment might be. A number of the physical and chemical variables can be deter-
mined on a one time analysis basis, however, where chemical release rates are
involved the estimate must be made from more than one analysis. In the case
of nutrients, these estimates may be made "directly" via time series analysis
using two different approaches: 1) i_n situ isolation of portions of the lake
bottom and overlying water suspected of contributing nutrients, or 2) incuba-
tion of sediment samples in the laboratory under simulated lake conditions
(i.e., time, temperature, oxygen content, and pH). In either case, sediment
nutrient release is measured as change in nutrient content of the overlying
water with time. The following is a brief discussion of factors considered
important to the description and analysis of phosphorus in sediments as a
plant nutrient source. Much of the same rationale could be applied to an
assessment of sedimented toxic materials.
Table 2 indicates considerable variation in anaerobic and aerobic phos-
phorus release rates for sediment from different lakes under J_n situ and
laboratory conditions. Anaerobic and aerobic phosphorus release rates are
both important since they represent conditions common to the two major sedi-
ment compartments of concern to the lake manager. These components are the
deep water (profundal) sediments (Mortimer, 1941, 1942; Livingston and Boykin,
1962; Delfino et al. , 1969) and the shallow water (littoral) sediments with
associated plant communities (Lie, 1979; Cooke et a!. , 1977; Barko and Smart,
1979). The profundal sediments may alternate between anaerobic and aerobic
conditions depending on lake stratification, while the littoral zone will
usually remain aerobic.
Holdren and Armstrong (1980) evaluated the effects of mixing, redox, tem-
perature, bioturbation, and sediment type on phosphorus release from profundal
sediment. They found that bioturbation by tubificids and emerging chironomid
larvae had the greatest effect on phosphorus release rates. This may be
associated with Lee's (1970) observation that the sediment mixing zone extends
5 to 10 cm below the sediment-water interface in lakes with a well defined
interface. In general, Holdren and Armstrong (1980) found that raising the
7
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Table 2. Measurements of phosphorus release (modified from Holdren and .Armstrong, 1980)
Sediment
Doboy Sound
Baldeggersee
Clear Lake 1
Clear Lake II
San Joaquin Delta
Lake Mendota
Lake Norriviken
Lake Trummen
Lake of Tunis
Lake Erie
Simulated Benthal
Deposit
Ursee
Lake Ontario
Fures
Lake Esrom
St. Gribso
Grane Langso
Lake Mendota
Buzzards Bay
Eel Pond
Lake Sodra Bargundasjon
Muddy River
Lake Warner
White Lake
White Lake
Lake Mendota
Lake Wingra
Lake Minocqua
Little John Lake
Experimental
System
Intact Core
In Situ
Laboratory3
Laboratory3
Laboratory
Laboratory
Laboratory3
Laboratory3
Laboratory3
In Situ
Laboratory
Laboratory
Intact Core
Intact Core
Intact Core
Intact Core
Intact Core
In Situ
In Situ
In Situ
Intact Core
Laboratory
Laboratory
Intact Core
In Situ
Intact Core
Intact Core
Intact Core
Intact Cone
Temp °C
27
27
27
27
11
22
10-15
4-10
7
7
7
7
7-13
1.5
20
8
20-30
20-30
2-23
4-21
3-18
4-16
Release Rate
Aerobic
0.031
-9.3
9.7
7.4
13
49
5
0.68
3
0.03-0.8
-2.0
-1.4
0.2
0.6
3.2
4.9-16.0
9.6
1.2
-1.9-33
-0.56-3.4
<0. 02-0. 37
<0. 02-1.1
mg m-2 day-1
Anaerobic
9-10
8.7
7.3
11
53
10
15b
25-30b
7.6
154
6-16
17.3
12.3
1.2
0.8
7-11
36
96
26
27.2
19.0
0.67-65C
0.95-2.9°
0.03-3.1C
0.02-3.8C
Source
Pomroy et al_. , 1965
Vollenweider, 1968
Porcella et al_. , 1970
Porcella et al_. , 1970
Porcella et a]_. , 1970
Porcella et aj!. , 1970
Ahlgren, 1972
Bjork, 1972
Bjork, 1972
Burns and Ross, 1972
Fillos and Molof, 1972
Tessenow, 1972
Kemp-Nielsen, 1974
Kemp-Nielsen, 1974
Kemp-Nielsen, 1974
Kemp-Nielsen, 1974
Kemp-Nielsen, 1974
Sonzogni , 1974
Rowe et a_[. , 1975
Rowe et al_. , 1975
Bengtsson, 1975
Fillos and Swanson, 1972
Fillos and Swanson, 1972
Freedman and Canal e,
Freedman and Canal e,
Holdren and Armstrong
Holdren and Armstrong
Holdren and Armstrong
Holdren and Armstrong
1977
1977
, 1980
, 1980
, 1980
, 1980
Columns containing dredged sediment and water.
Not designated as either aerobic or anaerobic.
c N2-treated cores with 2-3 mg I-1 dissolved oxygen in overlying water.
Complete citations in Holdren and Armstrong, 1980.
-------
temperature or lowering the oxygen concentration in overlying water also led
to higher rates of nutrient release. Suspending the sediment or lowering the
water temperature tended to remove phosphorus from the water column. They
also reported that sediment type affected the relative importance of different
incubation conditions. One example of this was that redox had more effect on
noncalcareous sediments while temperature had more effect on calcareous ones.
Most of Holdren and Armstrong's (1980) experiments ran for 10 days. Some were
continued for 20-30 days. In most cases, phosphorus equilibrium was reached
in 2 to 10 days. In some cases, the phosphorus levels increased slowly over
the whole incubation time. They concluded that, given the proper environ-
mental conditions, lake sediments can release significant amounts of phos-
phorus to overlying water, even when those waters are aerobic. Stauffer
(1980a) provides detailed comments and interpretation of Holdren and
Armstrong's results.
It is important to note that not all sediment phosphorus is exchangeable
or biologically available. Some forms are tightly bound to sediment or
actually incorporated as part of the structural matrix (Jones and Bowser,
1978). Therefore, total sediment phosphorus (bulk analysis) is of limited use
in determining the potential for nutrient release. The phosphorus most likely
to exchange readily with overlying water is that dissolved in the interstitial
zone (pore water). Because dissolved interstitial phosphorus is often present
at high concentrations (up to 5 mg/1 in Shagawa Lake, Minnesota according to
Larsen and Schults, 1978) in relation to the overlying water, there is a
strong gradient for diffusion of the interstitial phosphorus to the overlying
water. As this phosphorus moves to the water column it must be replaced by a
repartitioning of particulate phosphorus. Some partitioning forms undergo
transition more readily than others, thus they more readily replace liberated
interstitial phosphorus and are therefore more important in terms of internal
phosphorus recycling.
Williams et a]_. (1976) suggested that sediment phosphorus be divided into
three components based on chemical extraction techniques (Figure 2): "apatite
phosphorus" (A-P), "organic phosphorus" (0-P), and "nonapatite inorganic
phosphorus" (NAI-P). The NAI-P is particularly important in lake sediment/
water nutrient exchanges because it is thought to be strongly associated with
iron as a ferric-oxide-orthophosphate complex under aerobic conditions. The
iron complex dissociates under anaerobic conditions releasing phosphorus to
interstitial water or directly to the overlying water. For this reason,
Larsen and Schults (1978) have suggested that the amount of NAI-P in sediment
might be a more valid measure of the readily available sediment phosphorus
reservoir than either total phosphorus or pore water phosphorus. This is
substantiated by Williams and Mayer's (1972) conclusion that an increase in
apatite (phosphorus incorporated as part of the apatite crystalline structure)
with depth in Lake Ontario sediments was a product of diagenesis of other
forms of phosphorus to apatite with time. This transformation process and its
relationship to more readily exchanged phosphorus forms is shown in Figure 3.
Jones and Bowser (1978) indicate that processes of organic phosphorus minerali-
zation and the release of sorbed phosphorus under reducing conditions in lake
sediment can provide resolubilized phosphorus through diffusion from pore
water, but that this process will be offset by apatite formation. The process
may be responsible for the commonly observed phenomenon of decreasing quanti-
ties of soluble reactive phosphorus with increasing depth in lake sediments.
-------
(
IN No OH
I6H
85 C
_ EXTRACT
(
SAMPLE
0.22M No CITRATE /
OHM No BICARBONATE /
I.OM. No DITHIONITE (COB REAGENT)
I5MIN
85C
_ EXTRACT
RESIDUE
IN No OH
16 H
25C
__ EXTRACT
(^RESIDUE J
0.5N HCI
I6H
25C
EXTRACT
HCI04
DIGESTION
FOGG AND
WILKINSON
HARWOOD
etol
AFTER
CLARIFICATION
HARWOOD
etol
C ORTHOPHOSPHATE
IN COMBINED
EXTRACT)
HARWOOD
et al
WATANABE
AND
OLSEN
CORTHOPHOSPHATE\ /toRTHOPHOSPHATE\
EXTRACTED BY J I EXTRACTED BY )
NoOH REAGENT) y \j:DB REAGENT) ^/
Figure 2. Outline of analytical procedures used to determine apatite-P, NAI-P,
and organic P (modified from Williams et a]., 1976).
10
-------
CONCENTRATION IN OVEN-DRY SEDIMENT
ORGANIC
PHOSPHORUS
SORBED
ORTHOPHOSPHATE
TOTAL
PHOSPHORUS
Figure 3. Distribution of phosphorus in a sediment profile from Lake Erie
(from Williams and Mayer, 1972) hypothetical!;/ depicted and illus-
trating the diagenesis of mineral phosphate from amorphous to
crystalline form. Percentage of each form is given by area fields
(modified from Jones and Bowser, 1978).
One objective of sediment removal for nutrient control purposes might, there-
fore, be to skim the upper nutrient-rich layer of sediment off to reveal a
strata containing a less readily exchangeable nutrient form. Characterization
of pore water and solid phase sediment profiles from representative areas of
the lake will permit the development of maps to locate the horizontal and
vertical distribution of the relatively nutrient-poor strata. For phosphorus
limitation planning these maps should be based on the component phosphorus
contents of the sediments as proposed by Williams et a\_. (1976). Methods for
measuring these components are described by Sommers et a^. (1972) and by
Williams and Mayer (1972). Interstitial water extraction methods are
described by Holdren et aJL (1977). Sediment characterization data coupled
with time series temperature and oxygen profiles can be used to calculate the
potential rate of nutrient release from sediments.
11
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Determination of sedimentation rates of organic detritus and nutrient
elements is important in providing estimates of how rapidly sediment has
accumulated, or is accumulating, in the lake basin, and how rapidly the
effects of sediment removal might be expected to deteriorate if the sedimenta-
tion rate remains high. A common technique for estimating sedimentation rates
in the United States involves determining the ragweed pollen (Ambrosia) hori-
zon in sediments and measuring the depth of accumulation above that known
horizon (Bortleson and Lee, 1972; Craig, 1972). Another method for estimating
sediment accumulation rates employs the isotopes 210Pb (Koide £t al. , 1973)
and 137Cs (Pennington et al_. , 1973).
Jones and Bowser (1978) point out that pore fluid chemistry, solid phase
chemistry and sedimentation rate may be used to calculate the fluxes of
solutes (thus nutrients) into and from the sediments. Close examination of
sediment data in conjunction with physical-chemical data for the lake should
be useful to the lake manager in determining the significance of internal
nutrient release and thus the advisability of removing sediment to control
internal cycling. For example, if one has determined from the literature, j_n
situ, or laboratory studies that the phosphorus release rate from anaerobic
sediments is on the order of 5 to 10 mg m-2 day-1, and that 50 percent of the
lake basin in question is intermittently anaerobic for approximately 20 days
during the summer, then the potential phosphorus loading to the lake from this
source would be: 5 to 10 x 0.50 x 20 = 50 to 100 mg m-2 summer-1. If the
overlying water column was 10 m thick the resultant phosphorus loading to the
water might amount to 5 to 10 mg m-3 summer-1. A comparison of the above
phosphorus release rates with those of surface input to the lake and with
those of deeper sediment strata might assist in making a decision about sedi-
ment removal. For example, if it could be demonstrated that deeper sediment
released phosphorus at a much slower rate, say 2.5 to 5 mg m-2 day-1, the
loading to the overlying water column from the strata might be reduced to
approximately 2.5 to 5 mg m-3 summer-1, or about a 50 percent reduction. This
could translate to a similar magnitude reduction for chlorophyll a in the
lake.
A similar approach could be used to assess the impact of vegetated and
non-vegetated aerobic littoral sediments, taking into consideration the per-
centage of lake basin covered by each. In the case of vegetated areas (macro-
phytes) one should keep in mind that seasonal variation in nutrient release
may be critical (Barko and Smart, 1980; Carignan and Kalff, 1980).
The above mentioned analytical approach provides a simplistic example of
how a quantitative estimate of the effectiveness of sediment removal for
nutrient control might be compared to other major loading sources to the lake.
It also provides an idea of how long the effects (minus sedimentation rates)
might be expected to last. Not all of the phosphorus released from the sedi-
ment is recycled to overlying water, so the potential impact is probably over-
estimated. Stauffer and Lee (1973), however, have shown that significant
amounts of hypo!imnetically released phosphorus do reach the epilimnion in
shallow lakes as the thermocline is eroded by strong winds. One might improve
the accuracy of estimates for sediment-released phosphorus impacts on the lake
by applying Stauffer and Lee's mixing model approach, or by application of
Stauffer1s (1980b) "LAKETRANS" model which predicts thermal diffusivities and
12
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solute transport in stratified lakes. A thorough discussion of internal
nutrient loading in lakes and methods for assessing their significance was
completed recently by Stauffer (1980a).
3. Determination of Sediment Removal Depth
When deepening a lake to improve it for sailing, power boating and
associated activities, the deepening requirements are relatively straight-
forward. Projects aimed at controlling internal cycling of nutrients may
focus on open water area sediments or on littoral sediments and associated
macrophyte beds. Sediment removal requirements in these cases are not always
clearly defined.
Lake Trummen, Sweden, is perhaps the most thoroughly documented case
history of a lake subjected to sediment removal to control internal nutrient
cycling. Sediment removal depth in Lake Trummen was determined in a fashion
similar to the example given previously. Sediment was characterized chemi-
cally and physically, both horizontally and vertically, in the lake.
Digerfeldt (1971) as referenced by Bjork (1972) determined that sedimentation
rates in the lake increased dramatically during the period from 1940 to 1965.
About 40 cm of loose FeS-colored (black) fine sedimeni was deposited during
that time. Aerobic and anaerobic release rates of P04-P and NH4-N from the
black sediment layer and the underlying brown layer were shown to be markedly
different. Based on these differences a plan was developed to remove the
upper 40 cm of sediment. Results of the project are presented as case history
#1 in a later section of this paper.
Another approach to determining sediment removal depth has been used by
Stefan and Hanson (1979) and by Stefan and Ford (1975). Their approach was to
use a lake temperature model to predict the lake depth necessary to prevent
summer destratification. The model uses air temperature, dew point tempera-
ture, wind direction, solar radiation, and wind speed inputs at 12-hour time
step intervals, plus a consideration of lake morphology, to produce an output
of temperature prediction with depth. The approach used by Stefan and Hanson
(1979) is similar to that developed by Stauffer and Lee (1973) which shows how
windpower erodes the thermocline of northern temperate lakes. The major
difference is that Stefan and Hanson (1979) used their model output to predict
to what depth a lake must be dredged to maintain a stable summer stratifica-
tion, thus preventing enriched hypolimnetic waters from mixing into the epi-
1imnetic zone.
Use of the Stefan and Hanson (1979) model for determining sediment
removal depth assumes that a stable summer stratification is necessary to
prevent enriched hypolimnetic waters from mixing into the epilimnion. Based
on that assumption, Stefan and Hanson (1979) indicated that Hall Lake (one of
the Fairmont, Minnesota lakes) would require dredging to a maximum depth of
8.0 m. Other data presented by Stefan and Hanson (1979) indicate that dredg-
ing to 8.0 m may require the removal of 4.0 to 6.0 m of sediment. There is
no reference to the sediment volume involved. However, it would be sizable.
Hall Lake has a surface area of 2.25 km2 and a mean depth of 2.1 m. Lake
profile data presented by Stefan and Hanson indicate an accumulation of 2.0 to
10.0 m of sediment in the lake basin overlain with about 0.5 m of "boundary
13
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layer limnetic material". There is no apparent chemical or physical distinc-
tion between this layer and the deeper sediment. In fact, their core analyses
show relatively uniform phosphorus concentration in sediment from the surface
to a depth of 8.5 m (737 to 1412 mg kg-1 for 37 samples, with a mean of 1097).
There is no indication, however, of phosphorus release rates for sediment from
different depths. It is possible that the release rates from deeper sediment
might be less than that of the surface sediment. Furthermore, it is possible
that reduced nutrient release from the deeper sediment could be sufficient to
significantly reduce the adverse impact of nutrients on the overlying water
even though stratification was not stable (Bengtsson, et al. , 1975). If that
were the case, it argues for removal of much less sediment at significantly
less cost and perhaps fewer disposal problems. Therefore, it seems advisable
to examine phosphorus release rates for sediment from various depths before
adopting a lake temperature modeling approach to determine dredging depth.
Where sediment removal is contemplated to control nuisance macrophyte
growths and associated nutrient cycling, it is necessary to know what depth
must be attained to prevent the nuisance. Guidelines for determining this
variable are even less definitive than those for controlling nutrient cycling
from limnetic zones.
A number of factors contribute to the growth of macrophytes, e.g., sedi-
ment type, nutrient availability, temperature and light level. The Wisconsin
Department of Natural Resources presently uses light level (as determined by
Secchi disc determination of water clarity) as the chief determinant of
nuisance macrophyte growth. The method consists of determining the average
summer water clarity through periodic Secchi disc measurements, and then
estimating the maximum depth of macrophyte growth. The estimate is determined
from a graphic display of maximum plant growth depth (m) as a function of
Secchi disc (m). The straight line relationship is described by the equation:
Y = 0.83 + 1.22X
where, Y = maximum depth of plant growth in meters
X = average summer water clarity (Secchi disc) in meters.
The r2 for this relationship is 0.53, with N equal to 55. According to Dunst
(1980b) the relationship was derived from work done by Belonger (1969) and
Modlen (1970). Dredging projects designed for macrophyte control in Wisconsin
currently use this technique as a guide to prescribe dredging depths for
macrophyte control (Dunst, 1980b). However, Dunst (1980a) also indicates that
dredging depth need not always exceed the predicted Y-value to achieve effec-
tive rooted plant control. Simply increasing the water depth can result in
speciation changes, and in Wisconsin, nuisance surface growths are uncommon
where depth equals or exceeds three meters. Moreover, removal of soft organic
sediment to hard sand and clay bottom usually results in reduced plant growth
and density of problem species. Unfortunately, truly predictive methodologies
are undeveloped in these areas. It should be pointed out that the objective
of dredging for macrophyte control is not synonymous with macrophyte eradi-
cation. Any macrophyte dredging plan must take into consideration the pre-
servation of fish spawning and nursery areas, waterfowl feeding areas, and
other wild!ife habitats.
14
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4. Environmental Problems Associated with Sediment Removal
Peterson (1979) pointed out a number of potential environmental problems
associated with sediment removal. The problems generally can be classified
into in-lake and disposal site types. In-lake problems commonly center around
the resuspension of sediment during its removal when dredges, drag lines, and
clam shells are employed. One of the most common problems is the liberation
of nutrients with resuspended fine sediment. Phosphorus is of particular
concern due to its high concentration in interstitial waters in eutrophic
lakes and its affinity for finely divided particulate material. Dredge agita-
tion and wind action tend to move the disturbed nutrient-laden sediment into
the euphotic zone of the lake, producing the potential for algal blooms.
Churchill et a]_. (1975b) reported increased phosphorus concentration in Lake
Herman, South Dakota, coincident with dredging but they noted no increased
algal production. This was presumably due to the high turbidity level.
Dunst (1980a); on the other hand, found increased algal production in Lilly
Lake, Wisconsin when dredging began. This was a short term phenomenon
associated with the immediate time of active dredging. There were no signifi-
cant adverse effects.
Apparently, the reverse of increased algal production problems can also
be triggered by the resuspension of sediment. Reduced light penetration due
to turbidity will have a tendency to inhibit algal production. Lackey et al.
(1959) indicated this problem may be further aggravated by the tendency of
small phytoplankton to be adsorbed to the surface of precipitating fine sedi-
ment particles. A potentially more serious problem associated with finely
divided sediment in the water column is oxygen depletion. If the sediment is
highly organic, the particles quickly become bacteria-coated. The tremendous
surface area of these particles permits rapid decomposition and possibly
oxygen depletion. This factor may pose a problem in lake dredging projects
considering that the organic content of lake sediment may approach 80 percent
on a dry weight basis (Wetzel, 1975). To my knowledge there is no documenta-
tion of this problem in lake dredging projects, however, one should be aware
of the potential for it.
Another problem associated directly with resuspended sediments is the
liberation of toxic substances. Although small lake sediment removal projects
for the specific purpose of removing toxic substances are relatively uncommon,
the concern is becoming more evident (Matsubara, 1979; Sakakibara and Hayashi,
1979; Bremer, 1979). The major problem associated with toxicants and resus-
pended sediment centers around the fine particle sizes. Murakami and Takeishi
(1977) have shown that when dredging in PCB (polychlorinated biphenyl) con-
taminated sediments, up to 99.7 percent of the PCB materials adhere to sedi-
ment particles less than 74 u in diameter. This could be a particularly
perplexing problem in freshwater dredging due to the increased particle
settling times required for freshwater over those required for brackish or
marine waters.
/
A relatively common concern with lake dredging projects involves destruc-
tion of the benthic community. If the lake basin is dredged completely, two
to three years may be required to reestablish the benthic fauna (Carline and
Brynildson, 1977). However, if portions of the bottom are left undredged the
reestablishment may be almost immediate (Andersson et aj. , 1975) or within one
15
-------
to two years (Crumpton and Wilbur, 1974). In any case, the effect on the
benthic community appears to be of relatively short duration compared to the
longer term benefits derived from sediment removal. The same appears to be
true for game fish populations (Carline and Brynildson, 1977; Spitler, 1973).
The above concerns are associated primarily with dredging as a sediment
removal technique. The other major technique for sediment removal in lakes is
drawdown (lowering the water level) to expose the littoral sediments, or in
some cases (Born et aj. , 1973) the entire lake basin, followed by removal of
sediment with earth-moving equipment. Drawdown accompanied by bulldozer
sediment removal may pose additional nuisance problems such as noise, dust,
and truck traffic. It should be pointed out that hydraulic dredging might
produce an unplanned lake drawdown if the solid to liquid ratio of the dredge
slurry is low (typically 5% to 10%) and the dredge pumping rates are high.
5. Sediment Removal Methods
Once a lake has shoaled to the degree that it no longer serves a useful
lake-oriented recreational purpose, the methods available for amelioration of
the problem are limited. In theory, there are two approaches. One is con-
struction of an outlet control structure to raise the water level. This
approach has limited practical application due to problems of shoreline and
habitat destruction and flooding of private property. The second and most
common alternative is sediment removal.
The drawdown/excavate method of sediment removal was mentioned above.
The technique is summarized by Born et aj. (1973). It can be seen that the
applicability of this method is limited to lakes with control structures or to
those where pumping rates above water inflow rates can be achieved at reason-
able cost.
Dredging is a much more common approach to sediment removal. Pierce
(1970) described various types of hydraulic dredging equipment and provided
guidance on the engineering aspects of dredge selection, i.e., size of dredge,
pumping rate, pumping distances, etc. Barnard (1978) and Peterson (1979) have
described various grab, bucket, and clam shell dredges; hydraulic cutterhead
dredges; and specialized dredges to minimize secondary water quality impacts.
The conventional hydraulic cutterhead dredge (Turner and Fairweather,
1974) is the most commonly employed piece of dredging equipment. Hydraulic
dredges have a number of advantages over grab or bucket type dredges. They
have almost continuous operating cycles, allowing removal of large volumes of
material in a short time. Production rates in excess of 1,000 m-3 hr-1 are
not uncommon. These dredges can span virtually any horizontal reach of water,
due to their floating operations platform and, in general, have the capability
to be "environmentally cleaner" than grabs and buckets, due to their closed
cycle system of operation. This may not always be the case, however, because
by virtue of their ability to move large volumes of sediment, they also might
create large turbidity plumes, with resuspension of sedimented nutrients and
hazardous materials. The turbidity around the cutter of a hydraulic cutter-
head dredge increases exponentially with the thickness of the cut, rate of
swing, and cutter rotation rate (Barnard, 1978). The latter three functions
16
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*>
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RELATIVE PRODUCTION (PERCENT)
Figure 4. Relationship between the concentration of suspended solids 1 m from
the cutterhead and the relative production of a 61 cm cutterhead
dredge (from Barnard, 1978).
17
-------
are related closely to dredge production rate. Therefore, production rate is
closely related to turbidity around the dredge cutterhead. Using data from
Yagi et al. (1975), Barnard (1978) plotted the relationship of suspended
solids collected one meter from the cutterhead, to the production of a 61 cm
diameter cutterhead dredge (relative to its apparent maximum production rate).
He concluded (Figure 4) that it is possible to increase the dredge production
to the maximum rate (broken line in Figure 4), without generating excessive
turbidity levels. This involves optimization of cutting depth, swing rate,
and cutter rotation rate. Therefore, minimization of turbidity with a cutter-
head hydraulic dredge will depend as much on the skill and experience of the
operator as on the texture and particle size of the sediment and general
operating conditions.
Where minimizing resuspension of sediments is critical to the progress
and completion of a project (toxic substances removal), it may be necessary to
take special measures and perhaps to employ specialized dredging equipment.
Equipment types and capabilities, described by Barnard (1978) and Peterson
(1979), vary from modified grabs and innovative cutterhead design, the newest
being the disc bottom cutterhead (Breebot brochure), to air driven pumps. A
number of the specialized equipment variations have been used successfully in
removing toxic substances from Japanese waterways (Peterson and Randolph,
1977, 1978, 1979). Suda (1979) compared the suspended solids concentration
around a conventional cutterhead dredge to that around specialized Japanese
dredges. Results indicated an approximate 10-fold reduction in suspended
solids concentration around the specialized dredges. A serious drawback to
these results, however, is that no comparisons of production rates were made.
Recently, scientists at the U.S. Army Engineers Waterways Experiment
Station completed tests of the "pneuma" (air driven pump) dredge system.
While no mention was made of this system's ability to minimize resuspension of
sediments, it was indicated that the system was extremely inefficient because
of its high energy input requirements relative to dredge material production
(McNair, 1980). This may be indicative of the bid cost differentials between
hydraulic and specialized dredging equipment reported by Peterson (1979). In
view of this, one must examine closely how critical it might be to minimize
the resuspension of sediments in any sediment removal project.
6. Sediment Disposal Area
The major non-lake impact of sediment removal centers around the disposal
site. High population densities among a myriad of other things make it diffi-
cult to locate suitable sites. Another complicating factor is that the indis-
criminate filling of low-lying areas is no longer permitted under Federal Law
(Section 404, Public Law 92-500). Fill permits are required under Section 404
for virtually all projects where the lakes and any adjacent wetland proposed
for filling exceeds 4.0 ha (10 acres) in area. Guidelines for preparing
Section 404 permit applications are available from District Offices of the
U.S. Army Corps of Engineers.
In upland disposal areas (where Federal permits are not required) diking
operations are commonly employed. A major problem with upland disposal is
dike failure (Calhoun, 1979). Another problem not uncommon to dredging pro-
jects is underdesign of disposal area capacity. This problem is not always
18
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readily evident, but may become apparent as the project progresses. Disposal
of freshwater sediments may be particularly troublesome because of their
relatively slow settling rates (Wechler and Cogley, 1977). As disposal areas
fill, their ponding depth is reduced and so is their efficiency to retain
suspended solids. Disposal areas must be designed for the end of the project,
not for some intermediate point. Palermo et a_L (1978) have summarized impor-
tant technical information and research projects which will assist with the
proper design, construction, and maintenance of dredged material disposal
areas. Where standards for discharge from disposal areas cannot be met, it
may be necessary to treat the discharge. Barnard and Hand (1978) describe
when and how to treat if it becomes necessary. Other reports prepared by the
U.S. Army Corps of Engineers which are particularly useful in handling dis-
posal area problems include: Evaluation of Dredged Material Pollution Poten-
tial (Brannon, 1978), Confined Disposal Area Effluent and Leachate Control
(Chen et a\_. , 1978), Disposal Alternatives for Contaminated Dredged Material
(Gambrell e_t a_L , 1978), and Upland and Wetland Habitat Development with
Dredged Material: Ecological Considerations (Lunz et aj. , 1978).
7. Lake Conditions Most Suitable for Sediment Removal
Obviously, a lake with a filled basin might become a prime candidate for
sediment removal. How much sediment is removed will depend not only on the
amount of material obstructing use of the lake, but also upon the availability
of suitable sites for disposal. The latter need not necessarily be a deter-
rent to sediment removal. There are various productive use options available
for dredged sediment (Lunz et aj. , 1978; Walsh and Malkasian, 1978; Spaine et
aK , 1978). In the case of Nutting Lake, Massachusetts, where approximately
275,000 m3 of sediment were removed, the material is being sold for $1.00 m-3
effectively reducing dredged material removal cost to $0.45 m-3 (Manfredonia,
1980). This not only solves the disposal problem, but reduces significantly
the overall cost of the project
Due to the high cost of removing sediment from lakes, $1.00 to $15.70 m-3
as reported by Peterson (1979), it would seem that lake size might play an
important role in determining lake types suitable for dredging. That does not
appear to be the case. Surface area of lakes being dredged range from 2 ha to
over 1,050 ha, and volumes being removed range from a few hundred m3 to over 7
million m3 (Peterson, 1979). While technically there is no reason a large
lake cannot be dredged, there are some serious economic constraints. Most
projects, however, have not been scrutinized closely from the economic stand-
point.
Current information indicates that lakes with highly enriched surface
sediment, relative to underlying sediment, would benefit from dredging
(Andersson e_t al. , 1975; Bengtsson et a!. , 1975). Lake Trummen, Sweden showed
marked improvement when treated in this manner. Similar short term results
were observed when 25-50 cm of organic sediment was removed from Steinmetz
Lake, New York (Snow et aj. , 1980). This project differed significantly from
that at Lake Trummen, however, since clean sand replaced almost all of the
organic sediment removed (see case history #4 in this paper).
19
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Sediment removal will not be cost effective where nigh sedimentation
rates prevail. Sediment removal projects should have reasonable assurance of
extended longevity. Generally, the smaller the ratio of watershed area to
lake surface area, the greater the assurance of a lasting effect due to sedi-
ment removal. Watershed management practices should be encouraged with sedi-
ment removal projects where practical. However, massive watershed management
programs to protect small lakes could be counter productive. Uttormark et a1.
(1974) have pointed out that where watershed area to lake surface area ratios
greatly exceed 10 to 1, significant nutrient reduction from the watershed may
be impractical.
Generally speaking, small lakes with organically rich sediment, low sedi-
mentation rates and long hydraulic residence times would be ideal sediment
removal candidates.
IV. Case Histories of Sediment Removal
Sixty-four pre-implementation, on-going, or completed sediment removal
projects in the United States are summarized in the Appendix table. The table
identifies sediment removal projects in 22 states. There are no doubt others
which have not been included. Many lake sediment removal projects are under-
taken by municipal districts, counties and even state organizations with
little or no documentation of their effectiveness. The summary attests to
this by the large number of data gaps. In most instances, where data do
exist, they are not readily available. Published results are relatively
scarce and documented long term (5 years or more) effects are simply non-
existent.
The five case histories presented were included for one or more of the
following reasons: 1) they are examples of different sediment removal
approaches, 2) they have different objectives, 3) they point out some
strengths and weaknesses of the technique, 4) their results have been rela-
tively well documented. The first, and perhaps best example of a well docu-
mented sediment removal project is located in Sweden.
1. Lake Trummen, Sweden
Lake Trummen, Sweden is one of the most thoroughly documented dredging
projects in the world with both pre- and post-treatment information. This
project represents an attempt to reduce internal nutrient cycling by skimming
off a thin nutrient-rich sediment layer.
The lake has an area of 400 ha and, prior to restoration, a maximum depth
of 2 m and a mean depth of 1.1 m. By the early 1960s, domestic waste dis-
charge had deteriorated the lake to such a degree that the citizens of Vaxjo
considered filling the basin (Bjork et al. , 1971). Diversion of the major
pollutant sources had little effect on the lake. Sediment removal implementa-
tion in 1970 was preceded by two years of study and project plan development.
In 1970, one-half meter of finely divided organic sediment was dredged uni-
formly from the main lake basin (Bjork, 1974). In 1971, another half meter of
sediment was removed bringing the total volume to about 400,000 m3. This
20
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deepened the lake by approximately 40 cm. Dredged material was disposed of
partially in three macrophyte-infested, diked off bays. One large bay, repre-
senting approximatley one-third of the lake surface area and overgrown with
macrophytes, was preserved in total as a wildlife habitat. Dredged material
not placed in diked embayments was pumped to upland diked ponds where the
return flows were treated with aluminum sulfate to reduce the phosphorus
concentration in dredge slurry (M mg P liter-1) to about 30 ug liter-1.
Dried dredge material was sold as top soil for about $2 (U.S.) m-3. The total
phosphorus concentration in the lake prior to sediment removal was 600 pg P
liter-1. Following sediment removal the concentrations occasionally have
reached 70 to 110 ug P liter-1 (Bengtsson et al_. , 1975).
Bengtsson et al. (1975) indicated that phosphorus and nitrogen decreased
drastically (Figure 5) and that the role of the sediment in recycling nutri-
ents was minimized. The total and orthophosphate as well as the Kjeldahl-
nitrogen levels have remained significantly lower than before or during the
dredging program.
Biologically, the Shannon phytoplankton diversity index increased from
1.6 in 1968 to 3.0 in 1973 (Cronberg et aj_. , 1975). Secchi disc transparency
increased from 23 to 75 cm over the same period. Prerestoration mean annual
phytoplankton productivity was 370 g C m-2 (1968-1969), declining to 225 g C
m-2 following restoration (1972-1973). More than 60 percent of the annual
phytoplankton production in 1972-1973 was attributed to algae less than 10 urn
in size. Blue-green algal biomass was drastically reduced and some species
disappeared, notably Oscillatoria agardhii.
Before dredging in Lake Trummen, the benthic fauna was dominated by
oligochaetes and chironomids (Andersson et al_. , 1975). A year after dredging
tubificid oligochaetes and chironomids became much more numerous. The total
number of benthic organisms changed little. This was attributed to the
mobility of chironomid larvae and the fact that the dominant species swarmed
all summer, probably recolonizing newly dredged areas almost immediately.
The results at Lake Trummen are highly encouraging. Born (1979) indi-
cated that today Lake Trummen is not only a usable recreational resource, but
that the lake ecosystem has been restored. Figure 5 indicates that water
quality has been remarkably good for four years following dredging compared to
that prior to dredging. Jorgensen (1980), however, recently stated about Lake
Trummen that, "a deterioration of the water quality has been recorded during
the last 2-3 years (1976-1978), caused by the phosphorus input in the mixtures
of rain and wastewater discharged to the lake during heavy rainfalls." No
data were presented to determine what the nature of the deteriorated water
qua!ity might be.
2. Lake Herman, South Dakota
Lake Herman, South Dakota has a surface area of 546 ha, a maximum depth
of 2.4 m and a mean depth of 1.7 m. The lake has a 145 km2 agricultural
watershed. Row crop and small grain farming practices beginning in the early
1900s have resulted in the deposition of approximately 2 m of silt over the
entire lake basin. In 1970, 1971, and 1972, a 4.2 ha area of one bay in Lake
21
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LAKE HERMAN, SOUTH DAKOTA
Figure 6. Lake Herman, South Dakota (redrawn from Churchill et al., 1975a).
23
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Herman was dredged. This work was undertaken to test the effect of hydrauli-
cally dredging nutrient-laden inorganic silt. This project serves as an
example of a marginally successful, limited scale hydraulic dredging operation
to reduce availability of nutrients in a shallow, intermittently destratified
lake.
About 48,000 m3 of silt were removed from the dredge site, deepening it
from 1.7 m to approximately 3.4 m (Figure 6). Dredged material was deposited
in a 3.4 ha area adjacent to the lake (Churchill et a_L , 1975a). At the end
of the dredging project in 1972, drying had reduced the volume of the material
about 50 percent. No coagulants were added to the dredge slurry.
Orthophosphorus concentrations in the lake increased dramatically from
approximately 0.13 mg P liter-1 to > 0.56 mg P liter-1 shortly after dredging
began (Brashier et a!. , 1973). Phosphorus concentrations in the southeast bay
of the lake paralleled those of the main lake but appeared to lag behind by
several days (Figure 7). Despite the rapid increase in phosphorus concentra-
tions, they were not accompanied by major increases in phytoplankton produc-
tivity. This is not surprising, however, since Churchill e_t aj. (1975b)
indicated the lake was primarily nitrogen limited and that nitrogen frequently
declined to zero during algal blooms. Orthophosphorus concentrations in Lake
Herman prior to dredging were about 0.30 mg liter-1, during dredging (summer
1970, 1971, and 1972) they were almost uniformly 0.17 to 0.50 mg liter-1,
while after dredging (summer 1973, 1974, and 1975) they ranged form 0.10 to
0.50 mg liter-1 with a tendency toward the lower end of the range (Churchill
et aj. , 1975b). Churchill and associates were reluctant to conclude that
dredging was responsible for the increase in phosphorus concentration in Lake
Herman, however, they indicated that "there were no other noticeable environ-
mental changes that could readily account for this dramatic increase in phos-
phates". While no apparent increases in phytoplankton production resulted
from the increased phosphorus concentrations, the same may not have been true
if the lake was not normally nitrogen limited. If the large increases in
phosphorus were actually attributable to dredging one might conclude that this
particular dredging operation was inefficient in terms of limiting resuspen-
sion of sedimented material. While the project demonstrated that nutrient-
laden silt could be removed by hydraulic dredging, thereby reducing the amount
available for internal cycling in the area dredged, it left a question about
the efficiency of such operations.
3. Wisconsin Spring Ponds
The Wisconsin Spring Ponds (Sunshine and Krause Spring Ponds) Project
represents another of the more thoroughly documented records concerning the
ecological effects of dredging small lakes. The purpose of the dredging was
to deepen the ponds and increase their area to improve fish production. Inci-
dental to the deepening was the control of aquatic macrophytes. This was an
important plus for the project since Carline and Brynildson (1977) stated the
belief that aquatic macrophytes played an important, if not the most important
role, in determining the rate of filling in spring ponds. Presumably this
conclusion might have implications for other types of small closed basin lakes
such as Lilly Lake, Wisconsin (Wis. Dept. Nat. Resour. , 1969).
25
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The physical characteristics of the two spring ponds before and after
dredging are shown in Table 3. Sediment removal from the entire basin repre-
sents the most severe short run environmental impact possible. It completely
disrupts the benthic invertebrate population and removes almost all aquatic
macrophytes. In Sunshine Pond the predredging mean annual density of benthos
was 5,500 organisms m-2. Immediately after dredging the number was reduced to
84 m-2. Five years after dredging the density of benthic invertebrates had
increased to 16,000 m-2 in Sunshine Pond. Results for Krause Pond were simi-
lar. However, while most of the invertebrates prior to dredging were chiro-
nomid larvae, those five years later were primarily oligocheates which are not
an important trout food.
Carline and Brynildson (1977) indicated that after four years the impor-
tant fish food organisms, such as amphipods and chironomid larvae were still
in the process of reestablishing. This had a temporary, but statistically
insignificant, effect on fish production. Fish growth rates declined sharply
in 1971 during the dredging operation but by 1975 had returned to predredging
values.
Chara (a non-rooted, non-vascular aquatic macrophyte) covered 60 percent.
of the Sunshine Pond basin prior to dredging. Dredging removed essentially
all of the plant material. Five years after dredging Chara was reestablished
over 28 percent of the pond. However, its biomass was only about 10 percent
of the predredging level.
Lake dredging frequently is of concern to fisheries managers because of
the suspected adverse impacts on fish populations. Destruction of fish food
organisms is often the focal point of this concern. The above indicates that.
benthic invertebrate reductions had but a limited impact on the fishery in the
short run. Carline and Brynildson (1977) indicated that four to five years
after restoration by dredging the average density and biomass of fishable-size
fish were substantially greater than during the pre-dredging period.
4. Steinmetz Lake, New York
This case history is presented because of its relatively unique whole-
lake treatment technique, which consisted of complete drawdown, bulldozing out
of approximately 25 to 50 cm of organic sediment and replacement of the sedi-
ment with clean quarry sand. Macrophyte control and beach improvement were
the objectives.
Records indicate Steinmetz Lake, in Schenectady, New York was once a
quarry (Snow et aj. , 1980). The city acquired the lake and surrounding pro-
perty for a park in 1935. In 1936, the lake was drained and some deep holes
filled to improve swimming. It has been maintained by the city since that
time for recreational use.
The lake is 1.2 ha in area, with a mean depth of about 1.5m and a maxi-
mum depth of 2.1 m. It has a 28-40 ha naturally vegetated watershed which
serves as a public park (Snow et aj., 1980). Day use of the swimming and park
facilities declined rapidly since 1971 (over 17,000 users) to a low of approxi-
mately 6,000 in 1975. The decline was attributed to excessive weed growth and
26
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turbidity due to algal production, caused by nutrient loading from stormwater
drains and surface runoff, lawn fertilizing in the watershed, and the in-lake
macrophyte growth/death/decay cycle which was rapidly building up the organic
sediment (Snow et al. , 1980).
The restoration project was implemented in 1978. It included sediment
removal, diversion of stormwater drains, and the construction of a surface
runoff interceptor drains. Bulldozing 2026 m3 of sediment required much more
time than anticipated (six weeks) ,due to sinking of equipment in the soft
sediment (Snow et alI., 1980). Backfilling with clean sand was much less of a
problem. About 1843 m3 of sand was placed in eight days. This technique
results in no net changes of lake bathymetry, but produces a completely new,
clean substrate.
The short term results of the Steinmetz Project have been dramatic. Snow
et aj. (1980) and Bloomfield (1980) indicate the average summer Secchi disc
readings have improved from the 1976 and 1977 readings of 1.25 m and 1.33 m,
respectively to post-treatment readings in 1978 and 1979, of 1.47 m and 2.20
m, (maximum depth of the lake). Average summer chlorophyll a levels in 1976
and 1977 were 10.4 ug liter-1 and 5.8 ug liter-1, respectively. Post-
treatment levels have been reduced to 0.1 [jg .liter-1 in 1978 and 1.8 |jg
liter-1 in 1979. Aquatic macrophyte biomass, primarily Potamaogeton crispus,
ranged from 30-50 g wet weight m-2 before sediment removal and was reduced to
virtually zero after treatment. Snow et al. (1980) reported that after sedi-
ment removal P. crispus was observed growing at irregular intervals in
straight lines at a density of 2 to 20 plants m-2. Closer observation
revealed that the plants were growing where tracked vehicles had forced incom-
pletely removed organic sediment up through the sand cover during the sand
spreading operation. Snow et
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The fish population was dominated by slow growing yellow perch. Chemical
eradication and restocking with largemouth bass, sunfish and northern pike was
generally unsuccessful (Wisconsin DNR, 1969). Winter fish kills were common.
It was concluded that deeper water was required to manage for fish production.
The DNR recommended that at least 10 percent of the lake be dredged to a depth
of approximately 6.0 m. There is no indication of how that particular depth
or area was determined.
No action was taken on the DNR recommendation until 1976 when Wisconsin
received a grant from the U.S. Environmental Protection Agency for the purpose
of restoring Lilly Lake. The restoration plan submitted by the State called
for dredging about 665 x 103 m3 of sediment from the lake, deepening it to a
maximum depth of about 6 m. The objective of the project was to deepen the
lake sufficiently so that plant growth and winter fish kills would no longer
be serious problems (Wisconsin DNR, 1975). The project proposal called for
disposal of most of the dredge material in a nearby abandoned gravel pit.
Pre-implementation indications were that the sediment would flow to the
intake of a suction dredge and that cutterhead swing could be eliminated.
However, when dredging began in July 1978 it was quickly discovered that the
non-decomposed plant material in the sediment formed a tight matrix which
resisted flowing. From early July to October 26, 1978, approximately 360 x
103 m3 of sediment were pumped from the lake (Dunst and Beauheim, 1979).
During the 1978 dredging period in-lake 5-day BOD ran about 1-2 mg 02 liter-1
greater than during the previous year. Turbidity increased in the same manner
from 1 to 3 formazin units. Ammonia concentration increased dramatically from
approximately 0.01 mg liter-1 during most of 1977 to a high of nearly 5.5 mg
liter-1 when dredging ceased in October 1978. Although predredging records of
particulate phosphorus concentrations were incomplete it appeared that there
was a steady increase during dredging. Chlorophyll a levels in the lake
during 1977 were steady at about 2.5 to 3.0 |jg liter-1. In 1978 chlorophyll a
reached a high of approximately 27 |jg liter-1 immediately after dredging began
and then decreased to a level ranging from 12 to 18 (jg liter-1 for the
remainder of the dredging season. Dunst and Beauheim (1979) reported that
while productivity had averaged about 200 mg C m-3 day-1 during the summer of
1976 and 1977, it increased to an average of about 750 mg C m-3 day-1 in 1978.
Dredging commenced again in May 1979 and was completed by September
(Dunst, 1980b). Sediment removal actually exceeded 680 x 103 m3 and the lake
now has a maximum depth of 6.5 m. Additional Lilly Lake dredging data have
been reported by Dunst (1980c). The summer of 1980 was the first full growing
season since dredging. Dunst (1980a) has indicated that in-lake macrophyte
biomass of 200 to 300 g dry wt m-2 in 1977 was reduced to virtually zero when
dredging ceased in September 1979. Water quality since that time has remained
good and local sponsors are generally pleased with the outcome (Dunst and
Beauheim, 1980). It will be interesting to observe over the next few years if
macrophytes and internal nutrient cycling remain diminished as a result of the
lake deepening at Lilly Lake.
28
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V. Costs of Sediment Removal
Sediment removal costs vary widely depending on type and quantity of
sediment being removed; type of removal method, i.e., bulldozer, mechanical
hydraulic or special purpose dredge, sediment transport distance and method of
transporting; availability of disposal area; and ultimate use of dredged
material. Even within the same geographic region using the same dredging
equipment, costs may vary widely. Carline and Brynildson (1977) indicated
that cost per cubic meter of sediment removal to increase the volume of spring
ponds in Wisconsin ranged from $0.52 to $2.67. On a national basis the cost
differentials are much greater (Table 3). The Army Corps of Engineers has
shown a similar broad range in sediment removal costs on a national basis
(Saucier, 1976).
Generally speaking, per unit volume costs for sediment removal will be
inversely related to the total volume of material removed when disposal is not
a problem. Contaminated dredged material, requiring treatment and perhaps
upland containment, may increase cost by 3 to 5 times (Peterson, 1979). On
the other hand, sale of "desirable" dredged material may help offset the
removal cost. Nutting Lake, Massachusetts sediment is being sold for $1.40
m-3 (Manfredonia, 1980). A portion of the sediment from Lake Trummen, Sweden
was sold for nearly $2.00 m-3. Innovative disposal practices may signifi-
cantly reduce the cost of sediment removal. Closer examination of cost
figures is required before reliable generalizations can be stated.
VI. Summary
1. The purposes for removing sediment from a lake are 1) to deepen it
for improved boating, fishing, swimming and other recreational activities,
2) to prevent or reduce the internal sediment/water recycling of nutrients,
3) to reduce the effects of toxicants, and 4) to remove and reduce the growth
of nuisance aquatic plants. The expected results of these treatments are
improved dissolved oxygen concentrations; less sediment water nutrient
exchange, accompanied by reduced algal biomasses, increased clarity, and
reduced macrophyte nuisance problems. All of these goals may not be attained
in al1 cases.
2. The technique is effective for deepening, and when a lake has
shoaled extensively there may be no other practical method for restoration.
Sediment removal for deepening is accompanied by nutrient removal which no
doubt contributes to the overall improved lake quality in many sediment
removal projects.
The duration of deepening effectiveness will depend largely on lake
sedimentation rates. The one case history of skimming nutrient-rich surface
sediments for controlling internal nutrient cycling has been dramatically
successful for approximately six years. Signs of reduced water quality after
that time were associated with external nutrient loading, however, water
quality has remained better than it was prior to dredging.
29
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Macrophyte control by sediment removal is usually attempted through
deepening. The effectiveness of this approach is not clear at this point.
Sediment texture, nutrient content, and water temperature may be as important
as light level in determining if plants reinvade a dredged area.
3. The reduction of toxic problems in lakes by sediment removal is
technically feasible. This technique is being practiced at a high level of
sophistication in Japan. It requires special dredging equipment, confined
disposal areas and treatment of dredged material return flow waters. Costs
are usually double to quadruple those for removing uncontaminated sediment.
4. Relatively unsuccessful sediment removal projects can usually be
associated with extremely limited sediment removal (simply too small a scale
to be effective), an inefficient dredging operation (poor production of solids
in the dredge slurry, excessive downtime, extreme turbidity problems or inade-
quate disposal areas), or no attempt to control external loading of pollutants
to the lake.
5. Any project considered for sediment removal should be preceded by an
investigation of nutrient and sediment mass balances. Mass balance calcu-
lation will provide information on the relative proportion of pollutants
derived from the various sources, thus providing a mechanism to focus on the
critical compartments. Accurate portrayal of sedimentation rates permit the
extrapolation of sediment removal longevity, thus the comparative cost effec-
tiveness of sediment removal.
6. There are a number of environmental concerns associated with sedi-
ment removal activities. The majority of those associated with the sediment
removal phase (in-lake) appear to be of short duration and thus limited sig-
nificance when compared to the longer range effectiveness of the treatment.
This may not be the case when toxic substances are involved if sediment
removal is poorly managed. The effects of dredged material disposal may be
longer lasting. These effects will be tied directly to the method of disposal
and the quality of the dredged material. Filling of wetlands (no longer
commonly accepted) results in permanent destruction of that habitat. Enrich-
ment of agricultural land with non-toxic dredge material may have a long range
positive effect. Imaginative disposal plans can reduce the impact of disposal
and the overall cost of a dredging project.
7. The hydraulic dredge is the preferred piece of equipment for most
sediment removal operations. It is comparatively efficient in terms of pro-
duction and relatively clean (resuspension of sediments) in its operating
cycle. Special purpose dredges reduce resuspension of sediment, but are
relatively inefficient in terms of energy expended to remove a given volume of
sediment. Therefore, operational costs are usually higher than for conven-
tional hydraulic dredges. Their use may be warranted only when toxic sub-
stances are being removed.
8. An average unit cost for sediment removal cannot be stated. The
variables associated with dredging are so numerous that an average unit cost
figure becomes meaningless. About the only generalized statement that can be
made about unit cost is that it is likely to be inversely proportional to the
amount of sediment removed.
30
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9. There are still a number of uncertainties associated with dredging
as a lake restoration technique. Among these are the longevity of effective-
ness, the relative effectiveness of partial compared to whole-lake dredging,
how to dredge in order to minimize ecosystem shock and encourage rapid repopu-
lation of desirable organisms. In cases where nutrient control is of major
interest, more experience is needed with surface sediment skimming compared to
deepening as a restoration technique. A clearer picture of the comparative
effectiveness of these two techniques could have a major influence on future
projects. Costs could be reduced significantly if it was demonstrated that
surface skimming of sediments might accomplish the same objectives as deepen-
ing. More sustained research and monitoring of the effectiveness of diverse
sediment removal methods is required.
31
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41
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