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Lake and Reservoir Management 23:388-409, 2007© Copyright by the North American Lake Management Society 2007
Lake responses to long-term hypolimnetic withdrawal treatments
Gertrud K. NürnbergFreshwater Research
3421 Hwy 117, RR 1, Baysville, Ontario, P0B 1A0 Canada [email protected]
Abstract
Nürnberg, G.K. 2007. Lake responses to long-term hypolimnetic withdrawal treatments. Lake and Reserv. Manage. 23:388-409.
Hypolimnetic withdrawal is an in-lake restoration technique based on the selective discharge of bottom water to enhance the removal of nutrients and electro-chemically reduced substances that build up when the hypolimnion becomes anoxic. Comparison of water quality variables before and during treatment in about 40 European and 8 North American lakes indicates that hypolimnetic withdrawal is an efficient restoration technique in stratified lakes. Water quality improvement was apparent in decreased summer average epilimnetic phosphorus (P) and chlorophyll concentrations, increased Secchi disk transparency, and decreased hypolimnetic phosphorus concentration and anoxia. In particular, summer average phosphorus decreases were significantly correlated with annual water volumes and P masses withdrawn per lake area, indicating the importance of hydrology and timing of the treatment. Observations as well as models revealed that avoiding extreme temperature changes in the water column is critical for a successful application of the technique. The removal of colder bottom water may increase bottom water temperatures, which not only increases sediment release rates and sediment oxygen demand but, more important, may lead to thermal instability, resulting in enhanced entrainment of nutrient-rich hypolimnetic water and increased surface eutrophica-tion. Hypolimnetic withdrawal also improved water quality in man-made lakes with bottom outlets unless too much withdrawal led to thermal instability. It failed to have a positive effect in a shallow oligomictic lake, probably be-cause nutrient export was not much increased. A recognized disadvantage of hypolimnetic withdrawal is its impact on downstream waters, including eutrophication, temperature increase, oxygen depletion, and odor development. In the experiences evaluated, treatment of the withdrawal water ranged from no treatment in older remote applications in the European Alps, to passive treatment in wetlands and settling ponds, and modern waste water technologies in more recent applications. Overall, hypolimnetic withdrawal is an effective low-cost restoration technique to combat and potentially reverse eutrophication in stratified lakes and reservoirs.
Key words: eutrophication, hypolimnetic withdrawal, internal load, lake restoration, nutrients, phosphorus
In lakes with high internal phosphorus (P) loading from the surface layers of deep-water sediments, external load reduc-tion does not decrease the symptoms of eutrophication as expected, and levels of nutrients, phytoplankton, and hypoxia remain high (Sas 1989). To combat the high internal P load-ing and improve water quality of the lake, various in-lake restoration techniques have been applied. The ones used most in Europe and North America include the addition of chemicals (e.g., aluminum sulfate) to precipitate P, and the aeration or oxygenation of the hypolimnetic water to increase dissolved oxygen concentration (Cooke et al. 2005). The selection of such treatments can be facilitated by a decision support system as described by Schauser et al. (2003). These techniques attempt to decrease the trophic state by effectively
trapping P within the lake and increasing P sedimentation and retention or preventing P release altogether by main-taining oxic sediment surfaces. Their effectiveness is often limited due to undersized treatments because of high costs; furthermore, benefits may be short-lived because symptoms rather than causes are being addressed so that internal load rebounds upon discontinuation of the treatment, as in aera-tion. Even if these techniques are successful in decreasing the P concentration of the open water, they increase sediment P. Until oligotrophication dilutes the settling particles and P-poor sediments accumulate, these lakes cannot be consid-ered fully restored.
Lake responses to long-term hypolimnetic withdrawal treatments
389
A different approach is used by the lake restoration technique described here. Hypolimnetic withdrawal, as it is normally applied to a thermally-stratified lake or reservoir, replaces the surface or epilimnetic lake outflow with discharge of bottom water from the hypolimnion. Its timing is carefully planned (usually summer and early fall prior to turnover) so that it enhances the removal of unfavorable substances, such as nutrients and reduced chemicals from the hypolimnion, effectively decreasing P retention. The withdrawal water ideally is treated like wastewater to prevent contamination of downstream waters, although treatment has been variable throughout applications. The restoration technique of hypo-limnetic withdrawal was first reported in Poland (Olszewski 1973) and has been used frequently in Europe (Nürnberg 1987; Kucklentz and Hamm 1988; Hupfer and Scharf 2002; Schauser et al. 2003) and occasionally in North America (Nürnberg et al. 1987; Cooke et al. 2005). The overall positive assessment of the author’s previous evaluation of hypolimnetic withdrawal (Nürnberg 1987) has not changed because the benefits of this lake restoration technique have been supported by many subsequent applications.
The present study summarizes the effects of hypolimnetic withdrawal on the water quality of lakes, including more recent examples. Whenever possible, I first investigate the response of relevant variables for the periods before and throughout withdrawal in several case studies. Next, I investigate empirical relationships between these response variables with characteristics of the treatment. In this way I hope to summarize treatment effects and discern general patterns. Where appropriate, I discuss special characteristics and exceptions, often with help of observations and comments by the original investigators as expressed in the scientific literature. I also present model approaches that make it pos-sible to theoretically assess the probability of negative effects such as physical destabilization for future restoration projects. The application of hypolimnetic withdrawal in reservoirs and impoundments is discussed as well. Because the benefit to the lake has to be carefully evaluated with respect to potential adverse impacts on downstream systems, the effectiveness of various types of outflow treatments is also discussed.
Data collection and analysisA deliberate effort was made to review all hypolimnetic with-drawal applications in North America and Europe (Table 1). No hypolimnetic withdrawal treatments were found for other continents except for a few reservoirs with bottom outlets, but these studies were not included due to lack of information on water quality benefits. In several cases data published in the primary and secondary literature were amended with personal communication and unpublished data sets made generously available by the people responsible for the treatment.
Where possible, data were taken untransformed from the references; however, often variables had to be converted, summarized and adjusted so that the data analysis could be standardized. Usually, averages and medians for the strati-fied summer period (May or June to September, October or November) were used to evaluate withdrawal effects. In ad-dition, specific variables were computed that help determine the mechanism and effect of hypolimnetic withdrawal. One of these important variables is the summer bottom tempera-ture, an indicator of the potential changes in the heat budget of hypolimnia experiencing withdrawal. To obtain this vari-able, July and August averages of ~1 m above the bottom at the deepest spot of the lake (in the withdrawal basin, if there are several basins) were computed from temperature profiles. The variable fall turnover date was used as indicator of the duration of summer stratification. It was determined as the date of complete homothermy as apparent from temperature and other profile information. Where possible the anoxic fac-tor was determined from oxygen profiles and morphometric information according to equation (1) (Nürnberg 2004).
Anoxic Factor = (1)
where ti, the period of anoxia (days), ai, the corresponding area (m2), Ao, lake surface area (m2), corresponding to the average elevation for that period, and n, numbers of periods with different oxycline depths.
Much of the evaluation consists of differences in water qual-ity between the period before and during treatment. If a trend was apparent, latest available values were used to evaluate the changes from before and during treatment; otherwise (and in most cases) data were pooled and medians were compared. A variable, proportional change, was computed that expresses the change as proportion of the pretreatment values, according to equation (2).
Proportional change = (during − before) / before (2)
The amount of change is not always indicative of its statisti-cal significance, as Macdonald et al. (2004) noticed in Chain Lake, British Columbia.
Statistical analysis used in the evaluation is based on nonparametric statistics (e.g., medians and nonparametric confidence bands) because of fewer statistical constraints. In addition, Pearson regression analysis was used to compare the dependency of response variables, especially changes in water quality, to variables related to withdrawal.
Nürnberg
390
Tabl
e 1.
-Loc
atio
n, m
orph
omet
ry a
nd h
ydro
logy
of l
akes
with
hyp
olim
netic
with
draw
al, c
omm
ence
d in
yea
r of “
star
t.” R
ef =
refe
renc
es: 1
, Nür
nber
g 19
87; 2
, Sam
pl 1
992;
3, T
hale
r and
Ta
it 19
95; 4
, Ros
si a
nd P
rem
azzi
199
1; 5
, Špe
la R
ekar
, Env
ironm
enta
l Age
ncy
of th
e R
epub
lic o
f Slo
veni
a, p
ers.
com
m. ;
6, L
ivin
gsto
ne a
nd S
chan
z 19
94 a
nd P
ius
Nie
derh
ause
r, pe
rs.
com
m.;
7, H
upfe
r and
Sch
arf 2
002;
8, K
eto
et a
l. 20
04; 9
, Rön
icke
et a
l. 19
98; H
upfe
r et a
l. 20
00; 1
0, Z
aiss
198
4; 1
1, J
örg
Lew
ando
wsk
i, pe
rs. c
omm
. 12,
Sch
arf e
t al.
1992
; 13,
Ohl
e 19
72 c
ited
in K
lapp
er 1
992;
14,
Inke
Sch
ause
r, pe
rs. c
omm
; 15,
Ste
inbe
rg e
t al.
1982
; 16,
Kle
in a
nd C
horu
s 19
91; 1
7, D
unal
ska
2002
, 200
3; D
unal
ska
2007
; 18,
Ols
zew
ski 1
961;
19,
P
ette
rson
per
s. c
omm
.; 20
, Sos
iak
and
Trew
199
6; S
osia
k 20
02; 2
1, M
acdo
nald
et a
l. 20
04; 2
2, N
ürnb
erg
et a
l. 19
87; 2
3, H
enry
Run
ke, B
arr E
ngin
eerin
g C
ompa
ny, p
ers.
com
m.;
24,
Hol
t et a
l. 19
86; 2
5, L
athr
op e
t al.
2005
; 26,
KC
M 1
986.
M
orph
omet
ic P
ossi
- W
ater
Cou
ntry
,
O
bser
-
(O
sgoo
d)
bilit
y de
tent
ion
S
tate
or
S
tart
va
tion
Ad
Are
a V
olum
e z
z max
M
ixin
g In
dex
of
time
q sd
La
kea
Pro
vinc
e Lo
cb Y
ear
(aft
er)
ha
ha
m3 1
06 m
m
S
tate
c m
/km
m
erom
ixis
yr
m
/yr
Ref
Hec
ht
Aus
tria
E
A
1973
19
77
222
26.3
6.
43
24.4
56
.5
3 47
.6
2.49
2.
80
8.7
1K
lope
iner
A
ustr
ia
EA
19
75
1992
42
3 11
0.6
24.9
8 22
.6
48
3 21
.5
1.48
11
.50
1.97
1,
2K
raig
er
Aus
tria
E
A
1974
19
78
65
5.1
0.25
4.
8 10
2
21.3
0.
67
1.00
4.
8 1,
2L
eonh
arde
r A
ustr
ia
EA
19
80
1992
23
0 2.
3 0.
08
3.6
8
23.8
0.
65
long
c
2Pi
burg
er
Aus
tria
E
A
1970
20
00
264
13.4
1.
84
13.7
24
.6
3 37
.5
1.29
1.
90
7.2
1R
eith
er
Aus
tria
E
A
1976
1.
5 0.
07
4.5
8.15
36.2
0.
74
0.27
16
.5
1V
assa
cher
A
ustr
ia
EA
19
79
1992
12
0 4.
5
5.1
10.2
2
24.3
0.
70
c
2de
Pal
adru
Fr
ance
E
A
1976
4 80
0 39
0.0
97.0
0 25
.0
35
1 12
.7
0.79
4.
00
6.25
1
Kl.M
ontig
gler
It
aly
EA
19
79
1994
67
5.
0 0.
52
9.9
14.5
3
44.4
0.
97
10.0
0 0.
99
1,3
Var
ese
Ital
y E
A
1999
11 1
50
1,45
1.9
153.
65
10.6
26
1
2.8
0.42
1.
91
5.54
4
Ble
d Sl
oven
ia
EA
19
82
2000
75
4 14
3.8
25.6
9 17
.9
30.2
2
14.9
0.
87
1.50
11
.9c
1,5
Bur
gasc
hi
Switz
erla
nd
EA
19
77
1982
31
9 19
.2
2.48
12
.9
32
3 29
.5
1.53
1.
37
9.4
1L
ütze
l Sw
itzer
land
E
A
1982
20
03
602
12.8
0.
53
4.2
5.9
0 11
.7
0.31
0.
15
27.7
1,
6M
auen
197
4 Sw
itzer
land
E
A
1968
19
74
430
51.0
1.
99
3.9
6.8
1 5.
5 0.
25
0.60
6.
5 1
Mau
en 1
987
Switz
erla
nd
EA
19
68
1987
43
0 51
.0
1.99
3.
9 6.
8 1
5.5
0.25
0.
60
6.5
7W
iler
Switz
erla
nd
EA
19
62
26
3.
1 0.
33
10.0
20
.5
56
.4
1.54
1.
03
9.7c
1D
aem
man
&
Finl
and
EU
19
73
8
10 o
ther
lake
sA
rend
G
erm
any
EU
19
76
1985
2
980
514.
0 14
7.00
28
.6
49
2 12
.6
1.03
11
4 0.
25
9B
osta
l_D
am
Ger
man
y E
U
1979
19
80
1 26
0 11
9.0
7.45
6.
3 17
2
5.7
0.51
1.
50
4.17
10
Bur
g G
erm
any
EU
20
00
10.3
0.
50
4.5
26
3 13
.9
1.45
11
Gem
ünde
ner
Maa
r G
erm
any
EU
19
83
1989
43
7.
5 1.
33
17.7
39
3
64.6
2.
36
8.00
2.
21
1,12
Gre
bine
r G
erm
any
EU
19
68
13
Klu
ten
G
erm
any
EU
83
-85
25
0 8.
8
5.
1 0
0.
30
14M
edew
eger
G
erm
any
EU
89
, 91-
98
11
300
95
.4
10.1
5 10
.6
28.2
2
10.9
0.
90
0.62
17
.1
14M
eerf
elde
r M
aar
Ger
man
y E
U
1982
19
89
127
24.8
2.
27
9.2
18
3 18
.5
0.81
4.
50
2.04
1,
7O
bing
er
Ger
man
y E
A
1981
19
93
1 56
7 31
.2
2.18
7.
0 14
2
12.5
0.
59
0.27
25
.9
15Sc
hart
eise
n G
erm
any
EU
9.5
22.0
0
14Sc
hlac
hten
G
erm
any
EU
19
81
1990
42.0
1.
97
4.7
9.5
2 7.
2 0.
37
0.50
9.
38
16St
adts
ee B
ad
Ger
man
y E
U
1988
2 00
0 13
.7
6.88
6.
3 10
.5
2 17
.0
0.55
0.
20
31.0
14
W
alds
eeT
ress
ower
See
G
erm
any
EU
19
90
1
000
64.0
4.
90
7.7
20
2 9.
6 0.
71
2.26
3.
41
14K
orto
wsk
ie
Pola
nd
EU
19
76
86-8
9 10
2 89
.7
5.29
5.
9 17
.2
2 6.
2 0.
56
1.52
3.
87
17
1986
-89
Lake responses to long-term hypolimnetic withdrawal treatments
391
Tabl
e 1.
-Con
tinue
d
M
orph
omet
ic P
ossi
- W
ater
Cou
ntry
,
O
bser
-
(O
sgoo
d)
bilit
y de
tent
ion
S
tate
or
S
tart
va
tion
Ad
Are
a V
olum
e z
z max
M
ixin
g In
dex
of
time
q sd
La
kea
Pro
vinc
e Lo
cb Y
ear
(aft
er)
ha
ha
m3 1
06 m
m
S
tate
c m
/km
m
erom
ixis
yr
m
/yr
Ref
Kor
tow
skie
Po
land
E
U
1976
90
-94
102
89.7
5.
29
5.9
17.2
2
6.2
0.56
1.
52
3.87
17
19
90-9
4K
orto
wsk
ie 1
956
Pola
nd
EU
19
56
1956
10
2 90
.1
5.29
5.
9 17
.2
2 6.
2 0.
56
1.52
3.
87
1,18
Rud
nick
ie
Pola
nd
EU
160.
9
11
.9
0.33
17
Bru
nnsv
iken
Sw
eden
E
U
19T
reka
nten
Sw
eden
E
U
19Pi
ne
Alb
erta
N
A
1999
20
05
15 2
80
412.
0 20
.60
5.0
12
2 2.
5 0.
27
9.00
0.
56
20C
hain
B
C
NA
19
94
1999
44.0
2.
68
6.1
9 0
9.2
0.35
1.
75
3.49
c 21
War
amau
g C
N
NA
19
83
1985
3
700
286.
6 24
.76
8.6
12.8
2
5.1
0.31
0.
83
10.3
1,
22W
onon
scop
omuc
, C
N
NA
19
81
1985
59
9 24
.0
15.5
0 8.
5 15
.2
2 17
.4
0.69
4.
00
2.13
1,
22
shal
low
bas
inC
ryst
al
MN
N
A
Bef
ore
117.
0
8.
20
23
2001
Hay
es
MN
N
A
1984
24D
evil’
s W
I N
A
2002
20
05
686
150.
5 13
.9
9.2
14.3
2
7.5
0.41
1.8c
25B
allin
ger
WN
N
A
1982
19
91
1 17
2 40
.5
1.84
4.
5 10
1
7.1
0.40
0.
26
17.3
c 26
Ave
rage
1
943
121
17.2
9.
7 19
.1
2 18
.1
0.78
6.
6 8.
3 M
edia
n
43
0 42
2.
6 7.
0 14
.9
2 12
.6
0.65
1.
5 5.
9 n
51
32
39
34
35
38
33
35
37
32
32
a Fo
r tw
o la
kes
dist
inct
dif
fere
nt o
bser
vatio
n pe
riod
s w
ere
used
(M
auen
and
Kor
tow
skie
)
Seve
ral l
akes
are
kno
wn
to h
ave
been
trea
ted
besi
des
hypo
limne
tic w
ithdr
awal
: Rei
ther
See
was
dra
ined
and
fille
d w
ith c
lean
er w
ater
in 1
972,
then
an
iron
chl
orid
e tr
eatm
ent w
as c
omm
ence
d si
mul
tane
ousl
y w
ith h
ypol
imne
tic w
ithdr
awal
; Kle
iner
Mon
tiggl
er S
ee w
as o
xyge
nate
d un
der
ice
sinc
e 19
78, i
n Sc
hlac
hten
see
inflo
w P
was
pre
cipi
tate
d w
ith ir
on c
hlor
ide
sim
ulta
neou
sly
with
hy
polim
netic
with
draw
al, a
nd in
Cha
in L
ake
at le
ast t
hree
trea
tmen
ts h
ad b
een
appl
ied
befo
re (
dred
ging
, flus
hing
, aer
atio
n by
sm
all w
indm
ills)
.b
Loc
atio
n: E
A, E
urop
e –
alpi
ne; E
U, E
urop
e –
non-
alpi
ne; N
A, N
orth
Am
eric
ac
Mix
ing
Stat
e:
0 =
olig
o- o
r po
lym
ictic
, occ
asio
nally
str
atifi
ed in
the
sum
mer
1 =
Mon
omic
tic, s
trat
ified
thro
ugho
ut th
e su
mm
er s
easo
n on
ly2
= D
imic
tic, t
wo
stra
tifica
tion
peri
ods
per
year
, in
sum
mer
and
win
ter
3 =
Mer
omic
tic, a
cer
tain
laye
r at
the
deep
est d
epth
is s
trat
ified
all
year
long
d Su
rfac
e in
flow
was
enh
ance
d
Nürnberg
392
Lake characteristics and hypolimnetic withdrawal specificsHypolimnetic withdrawal has been applied as a lake resto-ration technique in Europe and North America (Figure 1). More than 50 applications are known to the author, mostly in central Europe and often in alpine regions. Many appli-cations were studied and reported in scientific journals and reports so that limnological data were available for about 40 lakes (Table 1), and withdrawal specifics (Table 2) as well as water quality trends were available for subsets of the total number of lakes. This technique has been used for about half a century, with many new installations materializing in the mid-1980s. A recent decline is apparent, with the exception of several new applications in North America (Table 1). Much of the original description was written in the language of the applying country (German, Italian, French, Slavonic, Pol-ish) and summaries and application instructions are mainly written in German (Klapper 1985; Hupfer and Scharf 2002). Nonetheless, several publications (see references in Table 1) and summaries about the technique have been published in English language journals (Nürnberg 1987; Nürnberg et al. 1987) and text books (Cooke et al. 2005). Despite these publications, this restoration technique continues to be unde-rused, especially considering it is low impact, inexpensive, and long-lasting. The present study hopes to encourage the application of hypolimnetic withdrawal in many more eutrophic lakes.
Treatment specificsThere are several ways of withdrawing water from the hy-polimnion, and various engineering constraints have led to different approaches. In particular, pipe diameter and length can be modified and are subjected to constraints. Many systems are based on passive siphoning (Table 2) because they are in remote areas like the European Alps or to limit energy costs (Chain Lake). The differential between the lake surface and the pipe outlet varies and creates a different head for the various applications. A large-diameter pipe may be less subjected to fouling via coating of the pipe interior but may require a valve to limit withdrawal during high water level if a passive system is employed (Kortowskie Lake). The pipe material of the first pipe was wood (Dunalska 2007) but newer applications usually use glassfiber and HDPE (high density polyethylene) pipes that have to be heavily weighted to keep them from floating (Lathrop et al. 2005; Al Sosiak, pers. comm., Alberta Environment, Calgary). The outlet structures may require various engineering solutions (Klapper 1985, Scharf et al. 1992, Macdonald et al. 2004). While almost all hypolimnetic withdrawal applications are conducted in lakes or reservoirs with outlets, the technique has recently been applied to a relatively large seepage lake
without an outlet (Devil’s Lake, Wisc.) using a siphon (Lathrop et al. 2005).
MorphometryHypolimnetic withdrawal has been used on lakes from 1.5 to 1,500 ha, spanning a 1,000-fold area and a 2,300-fold volume from 0.07 × 106 to 154 ×106 m3. Typically, only relatively small lakes are treated with hypolimnetic withdrawal (the surface area, Ao, of half of the study lakes is <44 ha). The size restriction is probably due to the limited amount of water that can be conveniently withdrawn through a pipe. To increase the withdrawal volume, several pipes have been installed in larger lakes, such as lakes Varese and Bred (3 pipes) and Pine Lake (2 pipes).
Figure 1.-Location of lakes with hypolimnetic withdrawal in Europe (upper panel) and North America (lower panel).
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Lakes are usually deep (the maximum depth, zmax, of half of the study lakes is >15 m, and the mean depth, z, is >7 m), considering their typically small areas. Accordingly, most lakes are stratified, mono- or dimictic, and 10 are even meromictic; only 3 lakes are polymictic or oligomictic (Table 1). Strong stratification is required because the prin-ciple of hypolimnetic withdrawal is to selectively withdraw
bottom water richer in nutrients (especially P) and reduced substances. Two shallow, less stratified, oligomictic lakes, Lützelsee (Livingstone and Schanz 1994) and Klutensee, stratified throughout the summer for periods of 1-2 weeks and occasionally exhibited anoxia throughout most of the water column before restoration. In a third shallow lake, Chain Lake, an artificial hole was created by digging 3 m
Table 2.-Hypolimnetic withdrawal specifics.
WD- Pipe Pipe Differ- TP_ Years Diameter Length ential zWD WD Rate qsWD QWD export Lake # Gravity cm m m m m3/sec m/yr 103 m3/yr kg/yr
Hecht 10 0 18 2 25.0 0.03 3.2 843 51Klopeiner 17 0 30.0 Kraiger 4 0 20 Leonharder 12 0 Piburger 30 0 8.9 639 11 23.0 0.011 2.1 284 7Reither 0 10 1 Vassacher 13 0 de Paladru 5 31.0 0.35 416Kl.Montiggler 15 1 14.0 0.2 10 2Varese 1 600 25.0 1.00 Bled 18 1 40 2 856 25.3 0.30 5.0 7 169 294Burgaschi 5 33 0.5 15.0 0.05 5.2 1 000 147Lützel 21 5.4 0.045 9.9 1 261 Mauen, 1974 6 0 30 0.5 6.5 0.067 2.0 1 000 617Mauen, 1987 20 0 30 0.5 6.5 0.067 2.0 1 000 Wiler 3 11 100 17.5 0.01 Arend 10 0 50 900 0.5 48.5 0.05 0.3 1 362 276Bostal_Dam 1 0 Burg 25.0 Gemündener 6 1 0 38.0 0.002 MaarMeerfelder 7 0 30 1.2 16.0 0.01 0.8 190 40 MaarObinger 12 1 41.5 500 0.4 13.0 0.07 Schlachten 9 Kortowskie 13 0 60 250 1 17.0 0.25 0.9 814 1986-1989Kortowskie 18 0 60 250 1 17.0 0.05 0.9 814 810 1990-1994Kortowskie 1 0 Square: 178 0.5 13.0 0.18 3.9 3 491 1956 40x50Rudnickie 20, 30 Pine 7 0 53 1 400 10.2 10.0 0.05 0.1 454 153Chain 7 0 46 170 5 6.2 0.05 1.1 501 52Waramaug 3 1 31.75 0 8.5 0.104 0.5 1 330 132Wononscopo- 5 1 0 15.1 0.015 0.8 201 22 muc, shallowCrystal 3 Hayes 100 50 6.0 0.45 Devil’s 4 0 51 1 676 2.2 14.1 0.15 0.3 438 352Ballinger 1 30.5 381 < 9.0 0.062 1.6 650 240
WD-Years: Years of operation corresponding to the observations
Gravity: 0 = passive siphoning; 1 = active pumping
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below the rest of the lake of 6 m depth (9 m deep, 70 m by 80 m area) and withdrawing nutrient-rich water from this location (Macdonald et al. 2004).
Since stratification and mixing potential depend on morpho-metric characteristics, I calculated 2 such indicators for the withdrawal lakes; the morphometric index (Osgood 1988) and the possibility of meromixis (Berger 1971). The mor-phometric index (equation 3, units = m km-1) is an indicator of the potential mixing of bottom with surface water:
Morphometric Index = z/Ao0.5 (3)
The larger the index value, the deeper the lake compared to its area. Therefore, large values represent lakes with strong and long stratification and potentially long anoxia, and small values represent lakes with more mixing, shorter stratifica-tion periods but higher oxygen depletion rates (Nürnberg 1995a). The morphometric index for most of the individual lakes, as well as the median and average for the whole data set, was >10 m km-1, indicating strong stratification (Table 1). The only lakes with values <3 m km-1, indicating early destratification and polymictic conditions, were multi-basin lakes (Pine and Varese). If the index is applied to the with-drawal basins separately, it is about 5 m km-1 (5.9 m km-1 in the South Basin of Pine Lake, Sosiak and Trew 1996; 4.95 m km-1 in the Gavirate Basin of Lake Varese, Chiaudani et al. 1995). The shallow lakes, Lützel and Chain, as discussed earlier, have relatively high indices, because of their small surface areas. Most lakes with hypolimnetic withdrawal exhibit fair to very strong stratification and hence are well selected for this treatment.
The possibility of meromixis (equation 4, units = m2) is computed as the ratio of maximum depth zmax and the double square root of the surface area:
Possibility of meromixis = zmax/Ao0.25 (4)
Meromixis is indicated when its value is >1 (Berger 1971). Furthermore, the positive deviation from one multiplied by zmax approximately indicates the thickness of the monimo-limnion.
Indeed, lakes described as meromictic in the original refer-ences have an index close to one or greater, except for the small volcanic kettle lake Meerfelder Maar (0.81); therefore this meromictic index is useful for determining meromixis in lakes where observations are not available. Knowing whether a lake is meromictic and at which depth meromixis commences is important because the depth between the monimolimnion and hypolimnion influences the location of the withdrawal outlet, as described below.
Hypolimnetic withdrawal depthMost of the withdrawal pipes were installed just above the bottom sediments of the lakes’ deepest regions where the highest concentration of noxious substances occurs. The dif-ference between maximum depth of the lake and withdrawal depth was usually not more than 1 m (Table 2; Figure 2), although in some cases it was not practical to put the pipe intake at the deepest location because it was far away from shore (e.g., Waramaug, Lac de Paladru). Two general excep-tions to this pattern are multi-basin and meromictic lakes. If the intake in multi-basin lakes was just above the bottom in a withdrawal basin shallower than the rest of the lake, the difference to maximum depth was higher (Bled, Pine, and Wononscopomuc). In meromictic lakes, withdrawal intake was located either at the lake bottom in the monimolimnion to reduce its volume (Kleiner Montiggler, Piburger, Meerfelder Maar, Burg, Gemündener Maar), or higher at the boundary between the monimolimnion and hypolimnion to prevent upward diffusion of nutrients and mixing of the deep water with hypolimnetic water, and ultimately the mixed layers of the lake (Burgäschi, Hecht, Klopeiner, and possibly Wiler). In Chain Lake an artificial hypolimnion was created by digging a 3-m deep hole into the bottom of the 6-m lake (Macdonald et al. 2004). The withdrawal outlet was situated just inside the dredged hole at 6.2 m depth to catch all P released from the anoxic sediments from this 9-m deep location.
Another exception to deep withdrawal depth was encountered in later treatment years in Lake Waramaug. Here the with-drawal depth was purposely raised to 6 m above the bottom
Figure 2.-Depth of withdrawal pipe position compared to maximum depth. Symbols indicate the mixing regime of the lake: 0 = polymictic or oligomictic; 1 = monomictic; 2 = dimictic; 3 = meromictic.
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sediments to avoid the discharge of highly polluted water resulting from a failed baffle system installed to precipitate and settle P and iron under vigorous aeration (described in a later section). The resultant nutrient and metal concentra-tions were above the state water quality guidelines, and unless outflow treatment could be improved, only elevating the intake could prevent its shut-down (Kortmann, pers. comm.). Raising the withdrawal depth drastically decreased the efficiency of the restoration, and therefore it finally was shut down in more recent years.
HydrologyA lake’s water budget, including temporal flow patterns as well as annual quantities, dictates how much water can be withdrawn from the hypolimnion in the summer and fall. Because the higher the withdrawal rate, the more nutrients and reduced substances are exported, high flows are directly related to the treatment’s potential benefits, especially when they happen in late summer and fall. Nonetheless, the treat-ment does not usually increase annual flushing to prevent upsetting the hydrological budget and decreasing lake level. Six exceptions are cases where the inflow has been augment-ed with high quality water, as in Bled, Leonharder, Wiler, Chain, Ballinger, and Devil’s Lake (Table 1). If such a water source is available, the potential efficiency of hypolimnetic withdrawal can be enhanced because more hypolimnetic water can be withdrawn than would be possible with an unchanged water budget.
In general, in lakes with hypolimnetic withdrawal, annual flushing rate (ρ, times per year) was high and water detention time (τ, years, which is the inverse of ρ) was low (median 1.5 yrs, Table 1). The annual water load, qs (m per year, also
Figure 3.-Annual water load (qs, open bars) and withdrawal water load (qsWD filled bars).
annual outflow volume, Q, per surface area, so that qs= Q/Ao or qs= z/τ), considers lake morphometry as well as hydrology and is a more complete description of a lake’s water budget. In half of the lakes, qs was relatively high, >5.9 m yr-1, which implies that a volume of water equivalent to a column of 5.9 m average height flows through the lake per year. The overall average qs was >8 m yr-1 (Table 1), and the lowest was 0.25 m yr-1 in Arend See, where treatment was installed because qs had been initially thought to be much higher (Hupfer and Scharf 2002).
The treatment specific withdrawal water load, qsWD, was lower in comparison because it represents the water volume withdrawn by the pipe only (QWD) instead of the whole lake outflow volume (so that qsWD= QWD/Ao) (Table 2; Figure 3). Both water loads are not necessarily correlated, and qsWD is expected to be a better predictor of treatment benefits than qs. While qsWD represents the observed withdrawn volume per year, the short-term withdrawal rate (Table 2) often presents a (maximum or average) design rate. Water is not always withdrawn at such a constant rate, and frequently water levels drop later in the summer causing the withdrawal rate to decline.
Effects of hypolimnetic withdrawal on treatment lakeThe most direct evaluation of a lake restoration treatment targeting eutrophication is the comparison of trophic state variables from before to during or after the treatment. Ac-cordingly, summer averages of phosphorus concentration, algal biomass, and Secchi disk transparency from before hypolimnetic withdrawal were compared with those during the treatment. After the trophic state evaluation, potential negative effects with respect to thermal stability were evalu-ated because hypolimnetic withdrawal may severely disturb a lake’s physical stability. To accomplish this task, first, any differences in thermocline depth and stratification duration between before and during treatment were noted. Next, changes in concentrations of dissolved oxygen and reduced compounds and extent of anoxia were determined. Further-more, observed and theoretical effects of temperature changes on oxygen depletion and P release are discussed.
PhosphorusHypolimnetic withdrawal treats a lake by preferentially exporting substances that accumulate under low redox con-ditions in the hypolimnion, such as phosphate and reduced metals and gases. For example, passive siphoning of bottom water increased TP export at least 2-fold in Chain Lake and Meerfelder Maar, 4-fold in Piburger Lake, and 38-fold on average in Devil’s Lake (data from studies cited in Table 1). Consequently, gradients across the thermocline became less
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Figure 4.-Epilimnetic (A) and hypolimnetic (B) P (µg/L) during hypolimnetic withdrawal versus before treatment (note the different scales on the x-axes). Regression line (in A, solid) is below 1:1 line (broken). Schlachtensee is not included in the regression, as it experienced simultaneous external load reduction.
pronounced, and the fertilizing effect of the hypolimnion to the trophogenic zone decreased. Therefore, TP decreases in the hypolimnion are expected to eventually improve water quality in the whole lake, including the epilimnion.
Indeed, P distribution within the treatment lakes changed drastically (Table 3). Before hypolimnetic withdrawal, sum-mer average TP concentration was up to 60 times higher in the hypolimnion than in the epilimnion. Once treatment began, TP concentrations of both layers decreased in most cases (Figure 4), and the differences between the layers grew smaller. As noticed before (Nürnberg 1987), the decrease was proportional to the initial P concentration, especially for the summer epilimnetic TP average, so that the higher the initial TP concentration, the higher the treatment effect. As a result, the log-log regression of summer epilimnetic TP averages
Figure 5.-Proportional change in epilimnetic P versus annual average water load due to withdrawal, qsWD. Regression line is shown with 2 significant outliers, Lake Ballinger (x) and shallow Lützelsee (out of range) removed. Zero change is indicated as broken line.
during treatment on those of before treatment was highly significant, even when the outlier value of Schlachtensee was removed (n = 20, R2 = 0.79, p = 0.0001; Figure 4A). Schlachtensee had a simultaneous treatment of the inflow with iron chloride precipitation that drastically decreased external P load. The regression line falls below the 1:1 line, indicating significant decreases in epilimnetic P concentra-tion (Slope is 0.74 ± 0.088 for SE, while intercept is not significantly different from zero).
If these observed changes in TP concentrations are caused by hypolimnetic withdrawal, they have to be correlated with variables that quantify this treatment. Because hypolimnetic withdrawal technique is based on withdrawing specific vol-umes of water, I analyzed the importance of hydrological conditions. Only lakes with a reasonable flow rate can be treated successfully. In Lake Arend the annual water deten-tion time, τ, was extremely large at more than 100 years, and the theoretical and observed effect of hypolimnetic withdrawal on TP concentration was negligible (Hupfer and Scharf 2002). The importance of the water flow was recognized by Scharf in Hupfer and Scharf (2002), whose experience with 2 meromictic lakes (Gemündener Maar, 9 yrs, and Meerfelder Maar, 4.5 yrs) indicated that τ should be ≤5 years to warrant enough water for hypolimnetic withdrawal to be effective. However, τ and its inverse, the flushing rate, are not significantly correlated with propor-tional epilimnetic TP change in this data set (n = 19, p = 0.73, after log-transformation of the hydrological variable), and neither is the annual areal water load, qs, (n = 20, p = 0.16, after log-transformation of qs). Instead, and as hypothesized above, the regression of proportional epilimnetic TP change
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Table 3.-Long-term averages of P concentration (summer stratification, maximum hypolimnetic and fall turnover) and average summer algal biomass indicators before and during hypolimnetic withdrawal.
Total Phosphorus (µg/L) Chlorophyll (µg/L) Secchi (m) Summer Hypolimnetic FallLake before during before during before during before during before during
Hecht 50 10 111 32 17.5
Klopeiner 15 12 239 165 5.2 6.0
Kraiger 15 14 2.0 2.8
Leonharder 41 34 2.0 3.0
Piburger 16 5.5 770 32 7.4 6.5 3 7.0 9.0
Vassacher 24 21
de Paladru 25 500
Kl.Montiggler 22 16 513 100 62 26 4 4.0 6.5
Bled 28 10 430 50 17 17 5.8 4.0 5.7
Burgaschi 80 20 300 60 50 20 0.3 1.5
Lützel 119 73 1 400 400 144 89 1.0 1.2
Mauen, 1974 100 40 2 000 500 70 50 1.4 1.4
Mauen, 1987 100 2 000 207 70 57 1.4 2.3
Wiler 1 800 500
Arend 163 141
Bostal_Dam 10 8 800 40
Burg 50 2
Gemündener 12 125 62 3.5 <3.5 Maar
Meerfelder 66 35 500 400 90 80 0.4 2.4 Maar
Obinger 118 65 500 34.5 33.3 1.2 1.3
Schlachten 600 30 2 500 180 800 30 60 8 0.5 3.0
Kortowskie 154 1.2 2.2 1986-89
Kortowskie 168 680 216 1.2 2.6 1990-94
Kortowski 1.2 1.7 1956
Pine 74 67.3 270 200 100 70 14.5 14 2.2 2.5
Chain 300 50 300 75 100 2.1 3.0
Waramaug 35 30 381 610 51 1.6 1.7
Wononsco- 25 12 473 89 5.6 5.3 pomuc, shallow
Devil’s 10 900 33
Ballinger 26 27 408 217.2 41 37 8 15 2.9
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Figure 6.-Proportional change in epilimnetic summer TP as function of TP exported during the corresponding period by hypolimnetic withdrawal. Regression line is shown. Zero change is indicated as broken line. Note that the export is an areal measure (total export divided by lake area).
on qsWD, which is based on the withdrawn volume instead of the entire outflow volume, is significant (n = 14, R2 = 0.35, p < 0.03), especially excluding outliers of Lake Ballinger and Lützelsee (n = 12, R2 = 0.72, p < 0.001; Figure 5). These relationships illustrate the importance of withdrawal volume rather than general hydrological conditions for the potential success of hypolimnetic withdrawal.
However, withdrawal volume is only a part of the treatment, since export is to be maximized by withdrawing at times when the TP concentration is highest, as in late summer (see section on Timing of hypolimnetic withdrawal below). Therefore, I hypothesize that the product of volume and concentration, (i.e., the TP export) is correlated to success. Indeed, the regression of epilimnetic change as proportion of the pretreatment concentration is significantly correlated with a logarithmic-transformed accumulated areal TP export (Figure 6; n = 12, R2 = 0.37, p < 0.05, and n = 11, R2 = 0.50, p < 0.01 excluding the influential outlier, Devil’s Lake). This result indicates that long-term continuous export decreases epilimnetic TP concentration. Nonetheless, it appears that duration alone does not always increase success. In contrast to Nürnberg (1987), the simple correlation of years of opera-tion (Table 1) with epilimnetic summer TP changes was no longer significant, although some of the same case studies were included in this analysis. One possible explanation is that the TP concentration decreases drastically at the begin-ning of the treatment and then slowly tapers off to a more continuous and slow decrease each year. This is evident in oligotrophic Piburger See (Prof. Roland Psenner, Innsbruck University, Austria, pers. comm.), eutrophic Lake Bled (Špela Rekar, Hydrometeorological Institute of Slovenia, pers. comm.), Obingersee (Schaumburg 1995), and other lakes (Nürnberg 1987). Furthermore, some operations are deteriorating for lack of maintenance, and raising of the pipe or fouling and narrowing of the interior have been reported (e.g., in Piburger See in 2004, Roland Psenner, University of Innsbruck, Austria, pers. comm.).
Hypolimnetic withdrawal effects on lake water quality can be predicted by a combination of P budget models and empirical regressions that relate water quality variables with P concentration (e.g., Nürnberg and LaZerte 2001). In particular, a P mass balance can reveal the importance of P loading from the sediments (Nürnberg 1998). As with any in-lake restoration technique, external P load has to be reduced first, or simultaneously, before any treatment of internal load promises sufficient decrease in lake P concentration. For example, a pretreatment study determined that the potential effect of hypolimnetic withdrawal would be marginal until external load was reduced in Lake Jabel, Germany by 75% (Kleeberg et al. 2001). In Pine Lake, efforts to reduce external load were made by implementing various best management practices throughout the watershed before and during the treatment. A detailed evaluation considering these loading
reductions and climatic variation promised a successful application of hypolimnetic withdrawal in Pine Lake for 7 of 10 years, whenever the water flow is sufficient (Table 1; Sosiak and Trew 1996).
In the study cases it is not always clear whether external load was constant during the course of the hypolimnetic withdrawal treatment. At least in one case, Schlachtensee, simultaneous external load reduction improved water qual-ity exceptionally fast. In the other cases reviewed here, no external load data can account for the overall observed water quality improvement during hypolimnetic withdrawal. Some of the variability in the observed P concentration during the treatment of the specific lakes may further be attribut-able to inter-annual climatic and hydrological variability. Despite this possibility, the correlation of the observed P decreases with accumulated areal TP export and qsWD, and its consistency, support the hypothesis that hypolimnetic withdrawal caused the observed improvement and, indeed, had a beneficial effect.
Dissolved reactive P
Most of the analysis concerning nutrients presented here is based on TP. Total P is a robust measure and less prone to analytical difficulties than other measures of P in water. Nonetheless, often a fraction called dissolved reactive P (DRP) is analyzed because it is largely biologically available and is taken up and used by plankton. Decreases in DRP can be expected to be more extensive than those found for TP as a consequence of hypolimnetic withdrawal, because P is released under anoxic conditions from the bottom sediments as ortho-phosphate. Field tests proved that P from anoxic hy-
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Figure 7.-Epilimnetic summer Secchi disk transparency during treatment versus before treatment. Most of the data points fall to the left of the 1:1 line, indicating increased transparency.
polimnia was typically up to 90% highly biological available and can be approximately determined as DRP (Nürnberg and Peters 1984; Nürnberg 1985). While there are less reliable data available for DRP than for TP, due to difficult and la-borious analysis of DRP in anoxic waters (Nürnberg 1984), several case studies indicate diminishing DRP concentrations with treatment, even if TP remained the same. In Lützelsee for example, DRP decreased much more than TP from values >700 µg/L to <20 µg/L at 5 m depth, which is about 1 m above the sediment (Livingstone and Schanz 1994). DRP decreased from 130 to 12 µg/L in the hypolimnion of Burgäschisee (Rene Gächter, EAWAG, pers. comm.) and from a median of 20 µg/L and 32 µg/L in the 2 pretreatment years to <10 µg/L in the summer epilimnion of the withdrawal basin of Pine Lake (Sosiak 2002); it was always less in the hypolimnion of the treatment basin as compared to the upstream control basin in Lake Kortowskie (Dunalska 2007). Decreased DRP concentration implies less available P for algal growth and hence indicates an improved trophic state in these lakes.
PhytoplanktonThe ultimate effect of a successful in-lake restoration tech-nique is the improvement of visual aspects of the water qual-ity, above all the reduction of algal biomass and blooms and the increase of water clarity, often measured as Secchi disk transparency or depth.
Summer average Secchi disk depth increased in 16 cases during withdrawal, in 9 of these cases by 0.9 m or more (Table 3; Figure 7). Secchi decreased in Lake Klopeiner (0.3 m) and Wononscopomuc (0.6 m), indicating decreased water transparency.
Changes in summer averages of Secchi transparency likely adequately indicate changes in algal biomass, often mea-sured as the algal pigment chlorophyll a, since it is unlikely that water color has changed during treatment. Secchi disk transparency represents both algal pigment and color in a wide range of different lakes and is correlated to a similar extent to both chlorophyll and color in summer average data for 200 worldwide lakes (Nürnberg and Shaw 1998). In the study lakes, Secchi has been more frequently determined than chlorophyll because it is a rather robust measure, easily obtained with minimal resources.
Nonetheless, chlorophyll data are available for a few lakes, mostly supporting the results of Secchi disk transparency. Summer average chlorophyll concentration decreased with the hypolimnetic withdrawal treatment in Lake Bled (from 12 to 5.5 µg/L), Piburgersee (from >6 to 3 µg/L) and Pine Lake (from 16 to 12 µg/L), but remained constant in Obingersee, and increased in Lake Ballinger (Table 3).
In addition to quantitative descriptions of diminishing algal biomass, qualitative statements are available indicating that the frequency of algal blooms declined during hypolimnetic withdrawal and that algal communities shifted to less noxious species (Kucklentz and Hamm 1988). Particularly noteworthy is the reduction of toxic Planktothrix rubescens blooms in at least 4 hypolimnetic-withdrawal lakes (Nürnberg et al. 2003). Planktothrix rubescens is a purple cyanobacterium that pro-duces pheopigments to catch light at lower irradiances and can release the toxin microcystin. Because the cyanophyte is adapted to cool temperatures and low light irradiance, it is typically located as a distinct layer at or just below the thermocline in the nutrient-rich metalimnion of hardwater, mesotrophic-to-eutrophic stratified lakes (Konopka 1982). Its buoyancy can be controlled with gas vesicles, hence supporting layering and preventing sinking. During mixing events in the colder part of the year, Planktothrix rubescens can rise to the top where it forms the typical red and then brownish colour of the water that is particularly striking under ice as it seeps into fissures and blowholes, creating dramatic patterns.
Gächter (1976) hypothesized that the bottom draw from the hypolimnion in the Swiss lake Mauensee was responsible for the eradication of Planktothrix rubescens by interrupting the metalimnetic flux of nutrients. Rigorous hypolimnetic withdrawal via 3 pipes (starting in 1980 to 1982) and 80% reduction of external phosphorus sources by sewage diver-sion (1982 to 1985) drastically decreased TP concentrations in Slovenic alpine Lake Bled (Vrhovsek et al. 1985). By 1990, the formerly abundant, toxic strain of Planktothrix rubescens had also declined and stayed low until 1999 when a manure collection system was installed close to the shores of the lake and thought responsible for increased nutrient concentration and Planktothrix rubescens abundance (Remec
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Rekar Špela, Hydrometeorological Institute of Slovenia, pers. comm.). In Lake Wononscopomuc a metalimnetic bloom was apparent in the years before withdrawal that was virtually absent in the withdrawal year 1982 (Kortmann et al. 1983). In the small German volcanic kettle lake Meerfelder Maar, Planktothrix agardhii, a closely related species, dominated the phytoplankton community until the second year of hy-polimnetic withdrawal operation, when it was replaced by a diverse diatom, dinoflagellate (Ceratium), and macrophyte community (Scharf et al. 1992). Another similar algae, toxic Oscillatoria limosa, disappeared in Kleiner Montiggler Lake after 9 years of hypolimnetic withdrawal and winter aeration (Thaler and Tait 1995).
In summary, it can be argued that hypolimnetic withdrawal increased water transparency in many lakes and reduced or eradicated toxic cyanobacteria like Planktothrix rubescens in several. In untreated lakes, algae were supported by entrainment and mixing of nutrients from the enriched bot-tom water. Planktothrix rubescens blooms occurred during periods of low light and enhanced mixing in several cases after applying whole-lake aeration and mixing as restora-tion treatment in three Swiss lakes (Bürgi and Stadelmann 2002) and several more European and North American lakes (Nürnberg et al. 2003).
Despite the general benefits of the hypolimnetic outlet, the flexibility of letting the water drain via a surface outlet is useful in lakes with occasional surface blooms (suggested by Ambühl, pers. comm. 1976) and applied in the Eau Galle Reservoir in Wisconsin (see Applications in reservoirs sec-tion).
Higher biotaThe influence of hypolimnetic withdrawal on higher biota has not often been examined. Theoretically, there should be a decreased loss of fish via outflow, since the withdrawal pipe often has a net on its intake to prevent courser mate-rial from entering and plugging the pipe. Such disruption of dispersion may not always be beneficial, especially when the downstream waters depend on the recruitment of fish from the upstream lake and vice-versa. Careful management for higher water levels and surface outflow at the times when migration naturally occurs may prevent the harmful effect of such a barrier.
Other more general effects on biota such as macrobenthos and zooplankton as well as higher vertebrates (fish and amphib-ians) can be expected to be positive because a decrease in trophic state from hyper-eutrophy or eutrophy to mesotrophy is accompanied by more beneficial benthic conditions and increased oxygen concentration. In Devil’s Lake, preliminary observations by scuba divers indicate that after 4 seasons of withdrawals the abundance of littoral freshwater snails have
decreased in response to a perceived decline in periphyton growth, thus preventing outbreaks of swimmers itch during the years of hypolimnetic withdrawal (Lathrop et al. 2005; R. Lathrop, Wisconsin DNR, pers. comm.).
Thermal stability and anoxiaThe potentially most severe negative effect of hypolimnetic withdrawal on a treated lake’s water quality concerns the physical limnology of a stratified lake. Because the treatment is based on preferentially withdrawing cold bottom water instead of warm surface water, it may increase hypolimnetic temperature. Such warming could cause the weakening of the lake’s thermal structure and destabilization leading to greater internal mixing. Furthermore, warming can negatively affect water quality as it increases oxygen consumption and P release from anoxic sediment surfaces. In the following sec-tion, I first present relevant variables from the case studies for years before and during treatment to determine whether and how often this potential negative side effect has occurred; and second, I investigate empirical relationships between these variables to determine their interdependence.
Temperature
There is no theoretical reason to anticipate any long-lasting changes in surface temperature due to hypolimnetic with-drawal (Dunalska 2007), and such changes are only reported in oligomictic Lützelsee (Livingstone and Schanz 1994). It was determined that hypolimnetic withdrawal accelerated the typical spring and early summer surface temperature increase by 20% according to heat budget calculations, or by 26% based on a comparison of observed temperatures of 2 pretreatment with 2 treatment years. The heat increase during spring was mainly due to short-lived temperature increases in the mixed surface layer and was not reported in other lakes. Surface temperature in the fall was more dependent on climatic conditions rather than hypolimnetic withdrawal, but hypolimnetic temperature increases were more evident.
Hypolimnetic temperature can be expected to increase in response to some removal of cold bottom water. The sum-mer bottom temperature averages in years before and during withdrawal, as measured within 1 m above bottom during July and August, were recorded (Table 4). Of the available temperature changes for 13 cases, 6 were positive, indicat-ing increased temperature, while the remainder was slightly negative or zero. Temperature increases were most conspicu-ous in the lakes Kortowskie (for all 3 periods), oligomictic Lützel and Mauen.
If thermal instability is induced by the treatment, more with-drawal should create more temperature changes. A direct measure of the extent of withdrawal is the withdrawal water load, qsWD, because it is based on annual average withdrawn
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Table 4.-Changes with respect to stratification (temperature) and oxygen depletion during hypolimnetic withdrawal.
Changes Fall Summer Depth of Turnover Bottom Depth of Period of Anoxic factor Thermocline Date Temperature Oxycline Anoxia before duringLake m days C° m days days/summer
Hecht 0 0 -5 0
Klopeiner 0
Leonharder -1.5
Piburger 2.0 -40 0 -3 -30
Vassacher 0
de Paladru 3.0 0 -3
Kl.Montiggler -90 0 -1 -90
Bled 14 -15 46
Burgaschi 0 -91 0 0 -91
Lützel 0 -27 3 0
Mauen, 1974 0 0 -2 -101
Mauen, 1987 1.7 -2.8 -163 0
Wiler -9
Meerfelder Maar 0 16 -1 -0.5 -107
Schlachten -2.5 64 26
Kortowskie 1986-89 -34 7 -120
Kortowskie 1990-94 0 1 4 -100
Kortowskie 1956 6.5 -108 11 -1 -56
Pine -1 17 16
Waramaug 0 -1.2 -0.7 -13 84 75
Wononscopomuc, -0.5 1 0.0 -2 -24 66 45 shallow
Ballinger 0.8 -1.5
volume per lake area. Indeed, the increase in bottom tempera-ture after beginning treatment is positively correlated with qsWD (n = 10, R2 = 0.53, p < 0.02; Figure 8) when the outlier values for all periods of Lake Kortowskie are removed.
As can be expected, in meromictic lakes bottom tempera-ture does not usually change with withdrawal, especially while meromixis is maintained, and significance is slightly increased when these lakes are excluded (n = 6, R2 = 0.86, p < 0.01). In general, close inspection of the data reveals that temperature does not change at qsWD below 1 m yr-1, but may increase at higher values. The largest increase, excepting the experimental periods of Lake Kortowskie, was in oligomictic Lützelsee (Table 4); it is not clear whether the increase is due to its mixing status or the high qsWD, since Lützelsee is the only oligomictic lake with relevant data.
Other measures of withdrawal, such as withdrawal rate and annual withdrawal volume, are not significantly correlated with temperature changes in this data set.
Because it appears that hypolimnetic withdrawal may indeed increase bottom temperature, the potential effects are inves-tigated in more detail in the following sections.
Stratification
Destabilization of the thermo-structure can negatively affect water quality and increase eutrophication if it occurs during the growing season when enhanced nutrient exchange across the thermocline would stimulate primary production. In lakes with extended stratification periods, however, induced mix-ing may increase the oxygen content before winter without increasing productivity and decrease or prevent anoxia and
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Figure 8.-Change in July-August average bottom temperature versus annual withdrawal water load. The regression line and mixing state according to Table 1 are indicated.
fish kills under ice, as attempted in Swiss Lake Burgäschi (H. Ambühl, EAWAG, 1976).
In many cases neither the thermocline depth nor the onset of fall turnover changed (Table 4). However, there are excep-tions. Several former meromictic lakes (e.g., Burgäschi and Kleiner Montiggler) turned dimictic because of the treatment; hence the stratification period was drastically decreased. In the first application of hypolimnetic withdrawal in 1956 in the south basin of Lake Kortowskie, Poland, early summer des-tratification was intentionally induced to shorten the period of anoxia (Olszewski 1973), which led to a deepening of the av-erage summer thermocline depth from 7 to 13.5 m (Table 4). Consequently, nutrient influx from the hypolimnion caused an algal bloom that was more severe than in pre-withdrawal years. Therefore, the operation was subsequently adjusted to prevent such early mixing, and the importance of operational details, such as flow rate and timing of withdrawal, have been investigated in Lake Kortowskie for almost 50 years (Mientki 1977; Dunalska 2003; Dunalska 2007).
Hypolimnetic anoxia
The effects of hypolimnetic withdrawal on anoxia were determined with the characteristics (1) depth of anoxia or oxycline, (2) duration of anoxia in days, and (3) the anoxic factor. In all 19 cases with available data, the depth at which anoxia started (oxycline) was lower (in 15 lakes) or remained constant (in 4 lakes) as compared to pretreatment conditions (Table 4). Similarly, the duration of anoxia during summer stratification decreased in 11 lakes, remained unchanged in one and increased in one (Lake Bled). Some of the shortened anoxic period may be due to shortened stratification, but the change in oxycline can only occasionally be explained by the deepening of the thermocline because it remained constant in many cases (Table 4). Where available, the anoxic factor, which combines both measures of spatial extent and period of anoxia, decreased in all lakes with available data (Table 4), as can be expected from these results.
Interdependence of hypolimnetic summer temperature and oxygen
The interdependence of changes in thermo-structure and oxy-gen concentration in the case studies was investigated with empirical regression analysis. If bottom summer temperature changes during the treatment had a net effect on anoxia, corresponding correlations should be significant. However, temperature changes are not significantly correlated with either depth of oxycline (n = 10, p = 0.81) or anoxic period (n = 10, p = 0.60). The lack of a correlation may be due to the limited range of temperature changes, indicating that hypolimnetic withdrawal in the case study lakes was carried out with minimal disturbances to the thermo structure. But it also may mean that a negative effect on oxygen by increased
temperature is compensated for by the beneficial effect of withdrawing reduced substances.
There are some exceptions. In Lake Kortowskie in the 1970s, increased hypolimnetic temperature led to increased oxygen consumption during and immediately after the stratification period, so that almost the whole water column became anoxic in the basin with withdrawal (Mientki 1977). In shallow Lützelsee, the warming of the bottom sediments and hypo-limnion led to more stagnant conditions and higher oxygen depletion rates (Livingstone and Schanz 1994); hence, the benefit of withdrawing water rich in reduced substances was curtailed.
Winter anoxia
While hypolimnetic anoxia during the summer causes phos-phorus release from the sediments and restricts fish and mac-robenthos habitat, anoxia under ice in the winter controls fish species richness and can cause winterkill (Nürnberg 1995b). Less winter anoxia may prevent fish kills, as was determined in 32 small Ontario lakes (Nürnberg 1995b). In that study, the potential of winterkill was quantified from the relation-ship (multiple regression) between fish species richness (or number of species) with winter anoxic factor (AFwin) and mean depth. If the model applies to small lakes in general, a threshold AFwin, above which winterkill is supposed to take place (i.e., fish species richness equals zero), can be deter-mined according to equation 5 (Nürnberg 2004):
Threshold of AFwin = −1 + 10(0.091 z + 0.804) (5)
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Figure 9.-Winter anoxic factor above which fish kill is predicted according to equation (5). Only lakes with a mean depth below 10 m that are not meromictic are shown.
The model may not be applicable for meromictic lakes that already have a stagnant anoxic bottom layer and is probably applicable only for lakes of <10 m mean depth.
In 21 treatment lakes with such characteristics, winterkill is predicted for an AFwin between 12 and 43 days ( Figure 9), which means an area equivalent to the whole lake surface area was overlain by anoxic water for that period of time. Winter anoxia under ice has decreased in Pine Lake (Sosiak 2002; predicted threshold AFwin = 17.2 d), Lake Kortowskie (Dunal-ska 2003; predicted threshold AFwin= 20.8 d), Piburger Lake (Pechlaner 1979; Psenner 1988; prediction not applicable because of meromixis) and Lützelsee (Gächter, pers. comm., predicted threshold AFwin= 14.3 d). All these lakes have a relatively low threshold indicating they have a tendency to winterkill. These lakes have continuous hypolimnetic with-drawal year-round, except for Pine Lake, where hypolimnetic withdrawal operates only during the summer months. These reductions in winter anoxia indicate the importance of hypo-limnetic withdrawal for the prevention of fish kill under ice, perhaps even when applied in summer only.
In Kleiner Montiggler Lake (Thaler and Tait 1995) oxygen-ation under ice was carried out to augment the beneficial effects of hypolimnetic withdrawal and prevent winter fish kill; therefore, in this lake increased oxygen under ice cannot be attributed to hypolimnetic withdrawal only. In addition, Kleiner Montiggler Lake was turned dimictic from mero-mixis, while also improving bottom water oxygen conditions via aeration during turnover events.
Reduced substances
Ammonium (NH4-) is a reduced gas that forms under reduced
conditions as soon as oxygen is depleted. It was reported to occasionally decrease during withdrawal. For example in Lake Waramaug, Connecticut, NH4
- mass was already lower in the first treatment year (1983), when NH4
- nitrogen was 3,600 kg in August and 4,300 kg in mid-September, as compared to a summer average >5,000 kg in the 3 preceding years and 7,500 kg in mid-September of 1980 (Nürnberg unpublished data).
Hydrogen sulfide (H2S) is an indicator of severe anoxia and low redox potential because it forms after the reduction of nitrates and metals. The concentration of this highly odorous and toxic gas decreased during the treatment in many lakes, even in those where no increases in oxygen were noted. For example, H2S decreased from 1 mg/L to 0.002 mg/L in the south basin of Pine Lake, Alberta (Sosiak 2002). It disappeared completely during the treatment in Kortowskie (Mientki 1977), and Burgäschi (Kucklentz and Hamm 1988), where it was abundant before hypolimnetic withdrawal. Hydrogen sulfide smells retreated from depths at 3 m be-fore treatment in Swiss Wiler See, to 6 m after 1 year and 20 m (i.e., 0.5 m above bottom) after 3 years of treatment (Eschmann 1969).
Prediction and maximization of benefitsModeling the effects of hypolimnetic temperature increasesThe comparison of water quality characteristics in the treat-ment lakes reveals that water quality improved in most cases. Nonetheless, under certain treatment conditions, water quality remained the same or worsened due to hypolimnetic withdrawal. In these cases, temperature increases were the most important trigger because they induced early mixing (e.g., Kortowskie Lake in 1956 and Lützelsee) and possibly increased anoxia and sediment phosphorus release; therefore, the quantification of these relationships by simple models was explored. In particular, 2 effects of hypolimnetic temperature increases were investigated in more detail: (A) shortening of the summer stratification period and (B) increased sedi-ment oxygen demand that may lower hypolimnetic oxygen concentration.
The shortening of the summer stratification period: A conse-quence of increased hypolimnetic temperature is predictable as the change (decrease) in fall turnover date by models based on summer bottom temperature for lakes with similar geography and morphometry, according to equation (6) (R2 = 0.70, n = 172, P < 0.0001 Nürnberg 1988b).
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Figure 10.-Observed fall turnover date change versus those predicted from bottom temperature changes for 17 years of hypolimnetic withdrawal in oligomictic Lützelsee.
Fall turnover date = 352 − 6.8 T (6)
Here, fall turnover date is expressed as Julian day, and T (°C) represents summer bottom temperature. After rewriting equa-tion (1), the fall turnover date change expected from a change in the average hypolimnetic temperature of any particular lake can be computed according to equation (7)
(Date2 − Date1) = 6.8 (T1 − T2) (7)
where 1 and 2 specify the values before and during the treat-ment. A negative value of the date difference means that fall turnover will happen earlier during withdrawal.
Increased sediment oxygen demand: The other expected consequence of increased hypolimnetic temperature is the in-crease of the sediment oxygen consumption rate or sediment oxygen demand (SOD). It is predictable according to equation (8) as revised from Livingstone and Schanz 1994),
SOD2/SOD1 = 0.068 x (T2 − T1) (8)
where SOD2 is expressed as the proportion of the value before treatment, SOD1.
To test these models I compared data for 5 pretreatment years with 17 treatment years from oligomictic Lützelsee that exhibited significant temperature increases in response to hypolimnetic withdrawal. In all observed treatment years the temperature changes were consistently positive, as the temperature rose from 13.2°C (average of 5 pretreatment years) to 16.4°C on average. During the same period, fall turnover was found to start on average almost one month (i.e., 27 days) earlier, moving from 4 October to 7 September. The corresponding fall turnover modeled according to equa-tion (7) predicted fall turnover to start on average 22 days early, which is close, but the individual years showed a high variation between predicted and observed fall turnover dates (Figure 10). It can be concluded that fall turnover changes are greatly influenced by bottom temperature changes. Long-term predictability is adequate, although individual years may not be predicted accurately, possibly due to climatic influences occurring after the July and August period. Hence the sum-mer temperature model can be used to predict any shorten-ing of the summer stratification period due to hypolimnetic temperature changes.
The SOD model of equation (8) predicts a change of SOD of 125% due to the temperature increase in Lützelsee. Si-multaneously, however, a decrease of oxygen demand in the water can be expected because of the withdrawal of reduced substances. Without a detailed heat budget it is not obvious which process is more important. In Lützelsee these effects appear to cancel each other, since inspection of the oxygen profiles reveals no long-term trend in either direction, and anoxia still rises as high as 3.5 m below surface, even after
almost 20 years of operation (Pius Niederhauser, EAWAG, pers. comm. 2004).
Increased sediment P release: Similar to the increased rates of oxygen demand, rates of nutrient release will also be higher at higher summer bottom temperature. Such an enhancement of areal P release rates by temperature has been documented in many lakes and experimental systems (Marsden 1989). Especially in shallow, polymictic lakes an increase in tem-perature may decrease the importance of the redox state (a measure of oxygen depletion and electron availability) of the overlying water because release was found to be indepen-dent of anoxia at temperatures above 17°C (Kamp-Nielsen 1975) and above 21°C (Holdren and Armstrong 1980). The apparent independence at high temperature can be explained by a severe anoxia at the sediment surface; Jensen and An-dersen (1992) found that the thickness of the oxidized layer decreases with temperature, and it was argued that at high temperatures SOD is so severe that P release from anoxic substances occurs directly into the oxic overlaying water (e.g., Boström et al. 1982)
The factor Q10, which quantifies the increase of a biologi-cal rate in response to a 10 degree temperature increase, is quite large as determined in laboratory studies: 3.5-7 in ex-periments purged with air (Jensen and Andersen 1992), 2.9 (standard error, SE, 0.5) in anoxic and 1.9 (SE 0.1) in oxic sediments of a hyper-eutrophic Finnish lake (Liikanen et al. 2002). This means that a bottom temperature increase due to hypolimnetic withdrawal of 3.2°C, as observed in Lützelsee (Table 4), may significantly increase P release. Simple core tube experiments (e.g., Nürnberg 1988a) could determine the effect distinct temperature increases are expected to have in specific applications.
Observations and model predictions indicate that a limited increase in bottom temperature may lead to enhanced anoxia
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and sediment P release during summer stratification and a shortened stratified period (because of decreased thermal stability). Such extended anoxia and increased P release (internal load) may be beneficial because it leads to enhanced withdrawal of reduced substances and nutrients, eventually decreasing sediment P content, as observed in Kortowskie Lake (Dunalska 2007). However, these increases in water phosphorus concentration may offset the immediate benefits to hypolimnetic water quality (i.e., may prevent immediate improvement of lake water quality). This delay should be acceptable as long as the water quality does not deteriorate be-cause of hypolimnetic withdrawal. The immediate expected benefit can be calculated from the increase in P export (or decrease in retention) as computed from the hypolimnetic P concentration before the treatment. The actual P export during hypolimnetic withdrawal is often higher than that predicted, probably because of enhanced anoxia and P release. The dif-ference between computed and observed P export can help to predict the long-term decrease in sediment P content.
Timing of hypolimnetic withdrawalThis investigation stresses the importance of the timing of withdrawal operation. Preferential operation of hypolimnetic withdrawal in late summer and fall during stratification is most beneficial for various reasons: (1) Export of nutrients and reduced substances that accumulate in the bottom water is maximized. (2) Changes in thermo and mixing status of the lake are minimized (bottom temperature and stratification).
If the amount of water available for withdrawal is limited to prevent extensive lake level decreases, the withdrawal rate should be managed to provide enough water later in the season, when hypolimnetic nutrient concentration is maxi-mal. For example, in Pine Lake a schedule of water level targets for specific dates throughout the growing season was established (Al Sosiak, Alberta Environment, pers. comm.). Based on long-term climatic conditions, there is enough water for withdrawal in 7 of 10 years. Withdrawing water later in the summer and early fall in dryer years would increase the benefit. In Devil’s Lake, there is only enough water for 1-2 months of outflow; therefore, hypolimnetic withdrawal is designed for September and early October only, the period with the highest hypolimnetic TP concentration (Lathrop et al. 2005). Late withdrawal also avoids or minimizes July-August temperature changes and the subsequent negative effects on thermo-structure and oxygen depletion.
Without management, the flow rate decreases at lower water levels in lakes with passive siphoning. The possibility of managing water flow in the summer by operating a valve was studied in Kortowskie Lake (Dunalska 2002, 2003; Dunalska 2007). In particular, the effect of differing daily withdrawal rates applied in different years to the same lake was systematically investigated. The same volume was
withdrawn (constant qsWD) at 2 different short-term rates: 4 years at the maximum rate of 0.25 m3/s and 5 years at a small rate of 0.05 m3/s (Table 2). The higher rate led to thermal instability throughout the summer and fall, while the lower rate did not. In particular, at higher rates, hypolimnetic tem-perature increased (for example, from 7-13°C in August) as well as oxygen depletion rates, which increased by 60% (for example from 0.12 to 0.19 mg/L/day at 10 m depth), and the stratification period was 5 weeks shorter. Conversely, during low flow operation oxygen depletion rates and stratification period were similar to those when no treatment took place, and withdrawal was sustained until the end of summer stratification.
Flow rate had an equally important effect on TP export in Kortowskie Lake (Dunalska 2002; Dunalska 2007). Model-ing the scenario of surface withdrawal, 42% of incoming TP would have been exported during the summer months (P retention was predicted as 58%). At the low flow rate in 1990-1994, TP export was 150% of input, and hypolimnetic withdrawal lasted the whole summer period from June-September. Conversely, at the high flow rate in 1986-1989 TP export was less, 115%, and 92% if the wet year 1987 is excluded, and lasted only from June-July because precipita-tion during summer was low.
In summary, if hypolimnetic withdrawal is too vigorous, early destratification could lead to algal blooms because growing conditions are better during high light conditions and warmer water temperature; it also shortens the period for the export of reduced substances and nutrients. Therefore, the timing of destratification is crucial and has to direct the operation of the treatment. In particular, timing and rates of withdrawal must be managed to prevent excessive hypolimnetic temperature increases.
Applications in reservoirsHypolimnetic withdrawal has been employed routinely in man-made lakes, as the outlet structures of many reservoirs operate at variable depths, facilitating hypolimnetic and metalimnetic withdrawal. The limnological effect on the reservoir’s water quality has not always been considered because the quality of the withdrawn water is usually more important, especially when it is used for human consumption. However, any adequate hypolimnetic withdrawal in eutrophic reservoirs can be expected to benefit its water quality, just as hypolimnetic withdrawal treatment does in lakes. Spe-cific design considerations including tables and formulas to determine pipe diameter and withdrawal flow rates for the bottom-withdrawal spillway in reservoirs were described by Pfost et al. (1993) and Rausch et al. (1993).
The effect of withdrawal depth on reservoir water quality was studied in medium-sized reservoirs in Wisconsin, U.S.,
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the Czech Republic and Germany (Lake Bostalsee, Zaiss 1981). In Eau Galle reservoir, Wisconsin (Ao = 0.6 km2; z = 3.2 m; zmax = 9 m; Barbiero et al. 1997), the normal operation using a hypolimnetic outlet was replaced by surface outflow in the growing season for 3 years. The intent was to decrease algal biomass by flushing the epilimnion and stabilizing the thermocline, hence preventing intrusion of hypolimnetic P across the thermocline. However, it was found that despite higher stability and lower hypolimnetic temperature, anoxia did not decrease, and hypolimnetic P increased as expected from the analysis presented above. External P loading was sufficient to keep epilimnetic algae well supplied with P, and flushing was not high enough to dilute biomass, so that there was no improvement with respect to surface algal blooms. In the Czech reservoir (Ao = 1.04 km2; z = 2.5 m; zmax = 5.5 m; Duras and Hejzlar 2001) using the bottom outlet meant more intense P release from sediments and lower P retention. Both aspects would eventually lead to lake improvements. Again, the phytoplankton biomass was not affected because it was not nutrient limited.
In contrast to Eau Galle reservoir, withdrawing surface instead of hypolimnetic water improved the water quality in man-made Twin Valley Lake, Wisconsin (Marshall et al. 2006). During bottom withdrawal operations flushing was so severe that stratification could not be established, and the im-poundment turned polymictic, increasing the flux of nutrients from the sediment into the trophogenic layer. Similarly, low water quality was blamed on bottom withdrawal in neighbor-ing reservoir White Mound Lake (Marshall et al. 2006).
In Cherry Creek Reservoir, Colorado, a deep water outlet withdraws water from the 1 and 2 m layer above the bottom. However, this reservoir is shallow and oligomictic, reduc-ing the efficiency of the treatment. Nonetheless, P retention was computed to be slightly lower than in surface outflow scenarios, and the beneficial effect on the reservoir’s water quality was modeled to be significant (LaZerte and Nürnberg unpublished studies, 2000). It is possible that anoxia and P release from the sediments increased because of elevated bottom temperatures, hence preventing short-term, but sup-porting long-term, decreases in nutrient concentration, as was speculated for Lützelsee (Livingstone and Schanz 1994).
Economically more important are the large reservoirs in the western U.S. of the Colorado system, including Lake Powell, Lake Mead, Lake Mohave and the Hells Canyon reservoirs on the Snake River. Metalimnetic and hypolimnetic outlets create complicated flow patterns that decrease P retention in the reservoirs. The most important nutrient source for Lake Mohave is nutrient-rich hypolimnetic water from Lake Mead via the Hoover Dam (Priscu 1982), so that hypolimnetic withdrawal benefits Lake Mead but decreases water quality in downstream Lake Mohave. In Brownlee Reservoir of the Hells Canyon Complex, hypolimnetic discharge leads to a
net export of biologically available P as DRP, soluble reac-tive P (Nürnberg 2002). The deep water discharge moves the thermocline deeper and increases the mixed epilimnion (to as much as 30 m depth in Brownlee Reservoir), leading to light limitation and a comparably low phytoplankton biomass in these deeply mixed water layers, despite avail-able P (DRP).
Occasionally, enhanced hypolimnetic withdrawal has been applied to reservoirs to increase their water quality. A short-term hypolimnetic withdrawal was employed in Lake Powell Reservoir on the Colorado River on the border of Utah and Arizona (Hueftle and Stevens 2001) to “freshen” the bottom water, which had become meromictic and oxygen deficient in previous years. Although the total withdrawal was only 2.8% of the full-pool volume and lasted only 9 days, it had a long-term beneficial effect on the water quality, disrupting the state of meromixis so that full lake circulation was rees-tablished during spring and fall, and decreasing hypolimnetic anoxia and conductivity (by diluting accumulated nutrients and redox dependant metals).
Effect on downstream water and treatment of outflow waterThe most controversial and problematic effect of the hy-polimnetic withdrawal is the impact on downstream water, including stream habitat. Not all of the downstream effects are unfavorable, however. In reservoirs, mid- and low-water withdrawals are used purposely (Kennedy 2005) to prevent warming of downstream waters, avoid organic substances from algae that would create odiferous problems in drink-ing water applications, and generally avoid the transfer of debris and other surface related substances being transferred downstream. However, oxygen decreases and increases of reduced substances such as ferrous iron and manganese as well as noxious gases such as hydrogen sulfide and methane, and the probability of eutrophication due to increased nutri-ent discharge may offset the above mentioned benefits if not dealt with adequately.
The most thorough treatment is based on chemical pre-cipitation and flocculation of nutrients and metals in a waste water treatment plant. In Gemündener Maar, withdrawal water was collected together with wastewater from public beaches to be treated in such a plant (Kucklentz and Hamm 1988); Leonharder outflow was treated in a communal sew-age treatment plant (Sampl 1992, 1993). Because the water is low in organic substances and particles, typically only a simple chemical treatment is necessary. In Ulmener Maar and Burgsee, treatment was provided by a movable flocculation tank called Pelicon (Phosphate Elimination Container; Keil 1995, Spieker 2002); this device was also employed to treat inflow water in several German lakes. In the former German
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Democratic Republic hypolimnetic withdrawal was often combined with irrigation projects in agricultural areas (Klap-per 1985), similar to a more recent application in Pine Lake, where water from one of the withdrawal pipes irrigates a golf course (references in Table 1). Other treatments include the construction or enhancement of wetlands (Pine and Chain) and settling ponds or baffle systems (Lake Waramaug and Wononscopomuc) to capture the nutrients and provide time for the equilibration with air, thus preventing major effects of nutrients and reduced substances on downstream waters (references in Table 1).
The prevention of air pollution by odor from hydrogen sulfide (H2S) is more difficult. An adaptation of the Pelicon apparatus that eliminates H2S from the water by iron sulfate chloride precipitation was used in Burg See (Hupfer and Scharf 2002), but was discontinued because of high costs after a year. Strip-ping and adsorption of gas in a specially constructed building with subsequent conveyance to an unpopulated area was used in Meerfelder Maar. Hydrogen peroxide (H2O2) addition to the outflow in Klopeiner Lake turned H2S into elementary sulfur, but this treatment was also discontinued because of expense (Kucklentz and Hamm 1988).
In many of the early applications the withdrawal water was not treated at all. If conditions became intolerable, the system was shut off in late summer and fall, such as in Hecht Lake and Lake Ballinger, hence drastically decreasing treatment efficiency. In some operations, pipe intake was elevated to prevent the discharge of low water quality water (Lake Wara-maug) or were finally discontinued permanently (Waramaug and Wononscopomuc), when outflow treatment needed to reach acceptable water quality standards as dictated by state laws were deemed too costly.
Downstream effects can be circumvented completely by discharging the withdrawal water after treatment back into the same lake. However, it is difficult to treat the water ef-fectively in little space and to prevent the warming of the water. In Lake Waramaug such a treatment was attempted unsuccessfully in a different basin from that of the regular hypolimnetic withdrawal, and hypolimnetic water was released back into the metalimnion after its treatment on a small peninsular by baffle systems and aeration of the natural occurring hypolimnetic iron (Nürnberg et al. 1987). Similar treatment was reported elsewhere, but no evaluation of its effect on lake water quality appears to be available (Hupfer and Scharf 2002; Cooke et al. 2005).
ConclusionThis analysis indicates that hypolimnetic withdrawal is an effective restoration technique in stratified lakes and deserves to be considered as a relatively low cost alternative to more invasive treatments. Improvement in water quality as de-
scribed by summer average P decreases were significantly correlated with withdrawal water load and hence require a certain flushing rate. Observations as well as models reveal that avoiding extreme temperature changes in the water column is critical for the success of the treatment. It is questionable whether hypolimnetic withdrawal is useful for shallow lakes unless temperature and stability can be maintained and nutrient export enhanced. In such lakes the addition of a chemical (like aluminum compounds) to ad-sorb and precipitate P may be more effective (Cooke et al. 2005). As with any in-lake restoration technique, external P load has to be reduced first or simultaneously, before any treatment of internal load promises a sufficient decrease in lake P concentration.
AcknowledgmentsThis study is based on the experience of researchers and lake managers dealing with hypolimnetic withdrawal as listed in Table 1. I highly appreciate the cooperation and sharing of information by these colleagues and their institutes in the form of internally published reports, data spreadsheets and personal communication. Furthermore, I want to thank Bruce LaZerte and my colleagues Dick Lathrop, Wisconsin, Al So-siak, Alberta, Roland Psenner, Austria, and Julita Dunalska, Poland, for providing inspiration, insight and criticisms.
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