Decomposition and nutrient dynamics of reed (Phragmites australis (Cav.) Trin. ex Steud.) litter in Lake Neusiedl, Austria

.4quaiic Botany, 43 ( 1992 ) 211-230

211

Else vier Science Publishers B.V., Amsterdam

Decomposition and nutrient dynamics of reed (Phragmites australis (Cav.) Trin. ex Steud. ) litter in Lake Neusiedl, Austria Peter Hietz Institute of Plant Physiology, University of Vienna, P.O. Box 285. 109! Vienna, Austria (Accepted 30 March ! 992 )

ABSTRACT Hietz, P., 1992. Decomposition and nutr;ent dynamics of reed ( Phragmites australis (Cav.) Trin. ex Steud. ) litter in Lake Neusiedl, Austria. Aquat. Bot., 43:2 i 1-230. Common reed leaf and stalk litters were exposed in coarse (5 mm) and fine ( 70/zm) mesh bags at five sites from the landward to the lakeward margin of the reed belt of Lake Neusiedl, Austria. Additionally, some samples were exposed above water and in a reedless pool. After 863 days, leaves exposed below water had lost 51-85% of the original ash-free dry weight, depending on site and hag mesh size. Weight losses of stalks and of litter exposed above water were considerably smaller ( 1532%). Decomposition was fastest at the iakeward site, slowest in the middle of the reed belt, and intermediate at the landward site. This rate correlated with concentrations of N, P, K and Ca, which were high after a short time in the litter at sites with higher decomposition rates. Leaves collected from one site and exposed at others showed that decomposition rates depended on site characteristics rather than on initial litter quality. Nutrient concentrations (N, P, K, Na. Ca and Mg), except of Fe, decreased during th¢ first month, and all, including Fe, increased subsequently until the end of the study period. Changes of cation concentrations can generally be explained by cation exchange processes. Photosynthesis ofepiphytes, carbonate precipitation and sediment incorporation can account for some of the changes observed. Increases in N and P, in contrast, seem to be controlled mainly by uptake by litter-settling micro-organisms. Of the various mathematical models tested, one including a temperature-independent labile fraction and a temperature-dependent recalcitrant fraction was the most successful.

INTRODUCTION

Although a steady decline of common reed (Phragmites australis (Cav.) Trin. ex Steud.) has been observed in other European lakes (see contributions to the special issue ofOstendorp, 1989), stands of the reed in Lake NeuCorrespondence to: P. Hietz, Institute o f Plant Physiology, University of Vienna, P.O. Box 285, ! 091, Vienna, Austria.

© 1992 Elsevier Science Publishers B.V. All rights reserved 0304-3770/92/$05.00

212

p. aiETZ

siedl have continuously increased during the past 200 years. In 1982 the reedbelt surrounding the open water covered 178 km: or more than half the total surface of the lake, and in some western and southern parts was more than 3 km wide. An extensive description of the limnology and geology of Lake Neusiedl is given by L~iffier (1979). In the la:~t few decades, eutrophication of the lake from fertilizers and runoff from adjacent, intensively cultivated felds and mainly domestic sewage, has increased. As the lake and its surroundings have become a favourite recreation area for nearby Vienna, this has put additional stress on the ecosystem. Whereas the bulk of nitrogen import is thought to be removed by denitrification in the reed stand (Van der Emde et al., 1986), phosphorus is a matter of permanent concern (Neuhuber et al., 1979). Because of their potential for controlling eutrophication, natural and seminatural wetlands have received much attention (Reddy and Smith, 1987). Cutting dead reed stalks in winter removes some nutrients from the lake. Summer cutting has been considered and would remove more nutrients, but may reduce reed growth by removing too many carbohydrates normally stored in the rhizomes. Additionally, reed is often burnt in winter, leading to a more homogeneous stand structure preferred by the reed cutters. These problems are discussed in more detail by various contributors in Aktionsgemeinschaft Gesamtkonzept Neusiedler See (1985). As only a small area of the reed-belt is cut every winter and primary consumption is negligible (Imhofand Burian, 1972), most of the stalks break down and decompose in the water. This study, therefore, has been undertaken to analyze the fate of the reed not harvested and to determine the role of decaying litter in the nutrient cycles of the reedswamp. SITES Five sites (3000, 2200, 1500, 800, and 100 m from open water, numbered 1-5, respectively) were selected along a transect from the landward to the lakeward margin of the reed stand, near Purbach in the northwestern part of Lake Neusiedl. Aboveground production of reed was studied by Sieghardt and Maier (1985). Suspended sediments are transported from the open lake by wave action and changing water level and deposited at Site 5, but no sedimentation was observed at sites more distant from the open water. The vegetation of emergent macrophytes is mainly dominated by Phragmites australis. At landward locations, with lower water levels, Cyperaceae and Juncaceae co-occur, and Typha angustifolia L. may also grow in deeper water. Lakeward Site 5 is dominated by monospecific stands of Phragmites, probably because other plants cannot tolerate the deep water (some 60 cm) and strong wave action. Oxygen concentr~,tions at Site 3 showed strong seasonal variations and at a depth of 20 cm, just above the litter layer, never exceeded 40% and were close

NUTRIENT DYNAMICS OF PHPwlGMITES LITTER

213

tO zero in winter. Redox potential in the sediment was shown to fall to 200 mV (Gunatilaka, 1985 ) and reed litter may be expected to decompose under mostly anaerobic conditions. Within the reed belt, areas, usually of a few to a few hundred square metres, occur, where no emergent plants grow. The origin of these pools may be natural or anthropogenic and has been discussed in some detail by Gunatilaka (1985) and Hietz ( 1991 ), but is still not fully understood. -

MATERIALSAND METHODS

Field procedures Dead leaves taken from the stalks, and stalks cut well above water, were collected on I l November 1987 and 2 November 1988. Subsequently, the material was air-dried in the laboratory for 1-2 weeks and subsamples were dried at 90°C to convert the results from air-dried to a dry matter basis. Airdried leaves (4.5-5.5 g) and stalks (7.5-8.5 g), cut to 10-cm pieces, were weighed to the nearest milligram and confined in plastic litterbags of l0 c m × 15 cm. Mesh sizes of 5 mm (coarse) or 7 0 # m (fine) were used to estimate the influence of the litter Fauna (Crossley and Hoglund, 1962). When !itterbags were exposed on 26 November 1987, a few bags were immediately returned to the laboratory to determine any losses caused by handling; none could be detected. In 1987, leaves and stalks in fine and coarse mesh-bags were exposed on the surface of the litter horizon in the reed stand, and leaves only in a reedless pool. Reed stalks may take a few years to break down and, thus, coarse and fine bags with leaves and stalks were suspended above water to follow the weight loss of standing dead material (Davis and Van der Valk, 1978). These samples were exposed at Site 3 in the centre of the reed belt, whereas only leaves in coarse bags were exposed at Sites l and 5. On 17 November 1988, leaves and stalks in coarse and fine bags were exposed at Site 3, and leaves in coarse bags at Sites 1-5. Leaves collected from Site 1 were additionally exposed at Sites 3 and 5, and vice versa. This translocation experiment was designed to show whether the different decay rates observed in Year 1 resulted from differences in site or material characteristics. Three to four litterbags of different series were bound together with a nylon thread and several of these were attached to a bar. They were placed randomly in an area of a few hundred square metres at each site, either below the water line so that they would never become dry, or suspended above water. Three bags of each litter-type were collected, first monthly, and later at longer intervals. The decomposition experiment ended on 7 April 1990, after 863 and 506 days for series 87 and 88, respectively. Unfortunately, several litterbags at Site 5, exposed in 1987, were lost, probably because they were

214

P.HIETZ

covered with sediment. Therefore, for this set only samples of Year 1 were available.

Laboratory procedures Litterbags exposed below water were thoroughly rinsed with tap water and briefly with deienized water to remove sediments, lake water and animals. Thereafter, the contents were dried at 90°C for at least 3 days, weighed, and the three samples were pooled and ground in a Cyclotec 1093 (Tecator AB, HSgan~is, Sweden) sample mill to pass a 0.5 mm sieve. Ash content was determined by combustion at 550°C for 4 h in a muffle furnace and decay calculated as ioss of ash-free dry weight (AFDW). Nutrients were extracted from 0.5-1 g samples with boiling 1 M HCI. Cations (Na, K, Mg, Ca, Fe) were analysed in a Perkin-Elmer 3030 (Perkin-Elmer, Llbedingen, Germany) atomic absorption spectrometer, and phosphate photometrically as phosphomolybdate complex. Kjeldahl-N was analysed after digestion with H2504 and a Se-K2SO4 catalyst in a Tecator Kjeltec System 1002 (Tecator AB, HiSgan~is, Sweden) distilling unit. All ash and nutrient measurements were repeated twice.

Statistics First analysis of variance (ANOVA) (as suggested by Wieder and Lang (1982) ) was applied to identify differences in the decomposition rates of the treatments (leaves vs. stalks; coarse vs. fine mesh bags; exposure above water vs. below water vs. in the reedless pool; exposure at Sites 1-5; litter originating from Sites 1 vs. Site 3 vs. Site 5). ANOVA produced suspiciously high significant differences for practically all treatments, even when they did not stand out graphically. As the time of exposure is a steady and independent variable and AFDW remaining can be described approximately as a function of time (see the discussion of models below), analysis of covariance (ANCOVA) with the time as a covariable was used. ANCOVA produced a far better resolution than ANOVA, and only ANCOVA results are discussed here. Data were checked for homogeneity of variance using the Fma~-test of Hartley and the Gma~-test of Cochran (Sachs, 1978 ), and for normality using the Kolmogoroff-Smirnoff test, or, for n < 51, the Shapiro-Wilk test (Zar, 1984). According to these tests, the data were normally distributed and variances were homogeneous. For both tests, arcsine transformed percentage values of AFDW remaining were used, as percentages lower than 30 and higher than 70 do not sufficiently satisfy normal distribution (Zar, 1984). Statistics were calculated using the GLM procedure of the Statistical Analysis Systems Institute (SAS) (1985).

NUTRIENT DYNAM ICS O F PllRAGMITES LITTER

215

RESULTS

Dry-weight loss Weight loss (Fig. 1 ) was rapid in the first month, accounting for approximately 10% of AFDW in leaves and 5% in stalks. Only the samples suspended above water lost weight at fairly constant and slow rates throughout the whole study period. Differences between coarse and fine mesh bags were significant ( P < 0.005, Table 1 ) after the first winter and tended to increase towards the end of the experiment. Decomposition in the reedless pool was slightly ( P < 0.05) faster than in the reed stand. Leaf litter in coarse bags decomposed fastest at Site 5 and slowest at Site 3. Decomposition at Sites 2 and 4 was faster than at Site 3 but slower than at Sites 1 and 5. In the translocation experiment (Fig. 2 ), the site of exposure had a highly significant (Table 2) effect on the decomposition rate, but the site of origin, which supposedly influenced the chemical composition of the leaves (Table 3 ), had no effect at all. Weight losses in both series were generally lower during the winter months.

Ash In submerged litter, ash content decreased steadily during Year 1 but remained constant thereafter. Final concentrations were about 1% dry weight in stalks and 2-5% in leaves at Sites 1-4. Only at Site 5 did non-combustible matter increase to some 20% in April 1990. Ash in samples suspended above water remained virtually constant for more than 2 years.

Nutrients Only the relative concentrations at the first sampling and at the end of the experiments are presented here (Tables 4 and 5). Nutrient contents of green leaves collected on I l August 1989 were used to estimate the nutrient changes of senescent leaves prior to leaf fall. Potassium content decreased strongly during the first month of exposure to about 5% of the initial concentration, except for material suspended above water. Subsequent accumulation was small and resulted i~na final concentration of about 10% of the initial; somewhat higher at Sites l ana 5 than at Site 3. Initial losses of Na were much lower than those of K, and by the end of the experiments concentrations were about 400-600 p.p.m., only a little lower than at the beginning. Calcium first decreased from 0.4-0.8% to 0.3-0.5% in leaves and from 750 to about 500 p.p.m, in stalks. Accumulation during exposure was low at Sites 3, 2 and 4, moderate at Site l, and high at Site 5. The

216

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1989

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Fig. I. Percent of ash-free dry weight (AFDW) remaining of the decomposition experiments started in November 1987. (a), Leaves and stalks exposed in the reedless pool and in the dense reed stand at Site 3; (b), leaves in coarse bags exposed at Sites 1, 3, and 5; (c), leaves and stalks exposed above water at Site 3. Vertical bars are S.E. t

217

NUTRIENTDYNAMICSOF PHRAG,~HTESLITTER TABLE 1 ANCOVA results of the liuer exposed at Site 3, 1987 series Source of variance Mesh size Litter tyl~e Mesh × litter Exposure Mesh × exposure Litter X exposure Mesh × litter× exposure Date Date × mesh Date X litter Date × mesh × litter Date×exposure Date × mesh × exposure Date × litter × expo~llre Date X mesh × litter X exposure Error

d.f.

SS

F

1 1 1 2 2 1 1 1 I 1 1 2 2 1 1

0.04546 !.29349 0.00202 2.20884 0.00169 0.24498 0.00000 3.83932 0.02435 0.11014 0.00384 0.01475 0.00162 0.0 i i 09 0.00013

8.51 242.14 0.38 206.75 0.16 45.86 0.00 718.73 4.56 20.62 0.72 1.32 0.15 2.08 0.03

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P 0.0038 0.0001 0.5389 0.0001 0.8537 0.0001 0.9857 0.0 0.0335 0.000 I 0.3967 0.2527 0.8593 0.1504 0.8743 MS: 0.00534

d.f., degrees of freedom; SS, sum of squares; MS, mean squares of the error.

high concentrations of Ca, Mg, and Fe at Site 5 in the first sampling of series 87 and the last sampling of series 88 resulted from the fact that these samples were only washed after having been dried, which did not remove the sediment to the same extent as when litter was washed before drying. Calcium, Mg, and Fe showed strong increases when inorganic sediment, consisting mainly of autochthonous protodolomite and Mg-calcite, was incorporated into the litter samples, whereas concentrations of Na, K, N, and P were little affected. Sediment deposited on the leaves could not be completely removed by rinsing the litterbags. This was not regarded as an unavoidable contamination, but rather as an intrinsic characteristic of litter decomposing at Site 5. Magnesium in leaves decreased from 1000-1500 p.p.m, to 600-700 p.p.m, at Site 1 and from 2300 p.p.m, to 900-1100 p.p.m, at Site 3, and remained about constant at 1700-2000 p.p.m, at Site 5. Subsequent increases were moderate and final concentrations were close to the initial ones, except for Site 5, where Mg increased to 0.6-0.9% dry weight after 1 year. Both Mg and Ca showed higher accumulations in the reedless pool at Site 3 than within the reed stand. Iron was not leached at all but accumulated continuously from 130 p.p.m, to 300600 p.p.m, at Sites 1-4 and to 2600-2900 p.p.m, at Site 5. From 5 to 10% of nitrogen was lost in the first month. Thereafter, N accumulated steadily in all samples and reached 2.6% in leaves at Site 1 and 2% at Site 3 at the end of series 87, without reaching a constant maximum after 29 months. Nitrogen increases were higher in summer than in winter, and

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Fig. 2. Percent of ash-free dry weight (AFDW) remaining of the decomposition experiments started in November 1988. (a), Leaves and stalks exposed in the water at Site 3; (b), leaves in coarse bags exposed at Sites I-5; l > 3, leaves collected from Site l, exposed at Site 3 etc. Vertical bars are S.E.

higher for litter in coarse than in fine mesh bags. Release and accumulation (to 0.4-0.55% dry weight) in stalks were moderate. Initial phosphorus losses (40-60%) were bigger than those of N. At the end of series 87, P had accumulated more than four-fold at Site l, and two- to three-fold at Sites 3 and 5. In the translocation experiment, concentrations of all elements depend only on the site of exposure, and not on their initial concentrations or differences

NUTRIENTDYNAMICSOF PHRAGAffTESLITTER

219

TABLE 2 ANCOVA results of the tran~location experiment Source of variance

d.f.

At site ~ From site m At site × from site Date Date X at site Date × from site Date × at site × from site Error

SS

F

P

2 2 4 1 2 2 4

0.20170 0.01562 0.01219 5.06 ! 37 0.02 ! 18 0.00333 0.01255

13.21 1.02 0.40 662.88 !.39 0.22 0.41

0.000 ! 0.3614 0.8091 0.0001 0.2523 0.8041 0.8004

190

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MS: 0.0"1635

~Litter exposed at Sites 1, 3 and 5. :Litter collected at Sites 1, 3 and 5 in November 1988. d.f., degrees of freedom; SS, sum of squares; MS, mean squares of the error. TABLE3 Initial mineral element contents of the litter used for the decomposition experiments, sampled in November 1987 and 1988 Ash N P K Na Ca Mg Fe (%drywt.) (%drywt.) (%drywt.) (p.p.m.) (p.p.m.) (p.p.m.) (p.p.m.) (p.p.m.) 1987

Site l, leaves 14.90 Site 3, leaves 12.80 Site 5, leaves 12.40 Site 3, stalks 6.30

!.07 0.91 !.27 0.41

0.083 0.051 0.112 0.045

3331 3641 5838 3848

836 919 429 1558

6466 8662 5758 429

1466 2384 1570 652

131 112 137 57

1.02 1.06 0.92 1.01 1.01 0.33

0.117 0.149 0.063 0.050 0.100 0.036

3676 5891 5496 5332 7226 5379

297 365 594 1076 527 1420

3542 6975 7530 8524 7644 735

1028 :.379 2190 2138 2298 586

124 119 167 147 126 67

1988

Site I,leaves Site 2, leaves Site 3, leaves Site 4,1eaves Site 5,1eaves Site 3, stalks

12.90 14.35 13.00 13.30 14.40 6.25

between materials from the three sites. Variations between leaves originating from different sites and exposed together were small and changes in concentration were more or less continuous.

Decomposition models Various mathematical models have been compared by Carpenter (1982), Wieder and Lang ( 1982 ), De Lyon et al. (1983) and Andr6n and Paustian (1987), among others. Only the single exponential (Olson, 1963; Table 6, eqn. (1)), double exponential (Wieder and Lang, 1982; eqn. (2)) and de-

!gg 84 166 235 42 173 250 5 4 120 0 55 70 176 296 176 284 514 47 653 1347

375 34 446 383 3 27 29 0 68 87 32 143 129 29 107 52 141 4119

5 coarse below

235 87 244

1 coarse below

Leaf litter

65 154 2182

138 26 77

67 24 [~2

66 2 66

284 3 9

594 62 306

276 108 212

3 coarse below

139 637

23 64

24 27

9 59

2 7

34 164

89 173

3 fine below

122 1511

30 148

32 !09

5 75

2 18

44 322

91 234

3 coarse pool

118 929

24 96

19 107

0 57

1 9

32 142

89 197

3 fine pool

148 447

64 42

61 52

42 30

56 30

138 286

105 152

3 coarse above

iValues are given as percentages of the values for litter collected in November 1987 (see Table 3).

N 11 August 21 December 7 April P 11 August 21 December 7 April K 11 August 21 December 7 April No I i August 21 December 7 April Ca 11 August 21 December 7 April Mg I 1 August 21 December 7 April Fe 11 August 21 December 7 April

Site Bag mesh size Above or below water or in pool

128 331

75 50

77 62

50 55

68 45

118 212

119 124

3 fine above

140 433

50 116

25 193

8 32

12 5

39 89

71 134

3 coarse below

Stalk litter

121 138

61 91

62 86

8 26

4 4

43 55

87 88

3 fine below

223 375

95 107

150 157

80 26

91 25

95 154

87 149

3 coarse above

151 365

99 106

149 154

95 36

106 29

111 86

127

3 fine above

Relative mineral element concentrationsof green reed leaves ( collected 11 August 1989) and of litter after 24 days exposure (21 December 1988 ) and after 863 days (7 April 1990), at the end of the decomposition experimentsI

TABLE 4

..e _~ .~

54 132 377

184 69 98

102 206

32 52

41 74

158 98 ! 53

8 16

347 6 19

12 91

122 473

266 81 184

82 22 173

92 211

3--, 1

250 85 177

1~ 1

131 280

25 48

36 74

15 97

5 12

60 364

86 206

54 1

148 288

106 172

84 68

64 199

9 10

40 71

89 157

1-,3

43 103 327

150 47 69

77 34 27

102 32 79

188 5 8

481 75 130

273 93 160

3--.3

113 303

39 60

31 32

47 94

6 5

80 79

99 149

5--.3

219 2038

199 894

120 1003

215 143

i0 14

43 100

90 177

1--,5

163 1721

83 474

59 505

77 71

4 !1

70 202

90 191

3~5

Values are given as percentages of the values for litter collected in November 1988 (see Table 3 ). "Translocations are given as the Site of collection-,Site of exposure. 3Leaf liner in fine mesh bag. 4Stalk litter in coarse mesh bag. SStalk litter in fine mesh bag.

I I August 21 December 7 April

Fe

I I August 21 December 7 April

Mg

! I August 21 December 7 April C,'/ I I August 21 E,'ecember 7 April

Na

N I 1 August 21 December 7 April P 11 August 21 December 7 April K I I August 21 December 7 April

Site translocation:

51 186 232 i

120 73 725

53 50 748

98 113 90

202 6 12

263 62 158

249 91 194

5--.5

68 124 321

187 45 71

89 36 48

55 22 142

191 5 10

207 38 43

250 91 162

2

39 109 456

105 52 108

41 29 40

110 36 108

268 7 12

456 86 352

245 94 166

4

96 184

44 57

34 21

59 83

5 4

56 84

88 125

33

112 126

73 134

73 i 22

16 48

6 4

8i 100

91 106

34

124 96

73 98

80 93

15 35

5 3

72 53

88 !12

3s

Relative mineral element concentrations of green reed leaves (collected 11 August i 989 ) and of litter after 34 days exposure (21 December 1989) and after 506 days (7 April 1990), at the end of the decomposition experiments ~

TABLE 5

222

P. HIETZ

'FABLE 6 Decomposition models to describe ash-free dry weight losses, evaluated by non-linear regression

(l) (2) (3) (4) (5)

W,/Wo=exp(-kxt) W,/Wo=aXexp( -k~ x t ) + ( 1 0 0 - a ) X e x p ( - k , x t ) Wt/Wo= exp(kl ) × (exp( - k 2 × t) - 1 ) It'd l~o=a × exp( -k~ × t) + ( 1 0 0 - a ) × exp( - k2 ×)qum' )

Wt/Wo=a ×exp( - k j × t ) + ( 1 0 0 - a ) ×exp( -k2×Qsu~.' )

Wo, initial weight; Wt, weight remaining at time t (years); k, kt, k2, decay constants; q, component decomposing with rate k~; Tsu,~',Qsum',temperature dependent factors.

caying coefficient (Godshalk and Wetzel, 1978; eqn. (3) ) models will be considered here. Few authors have included temperature in their models, although it obviously is an important factor controlling decomposition. Two temperaturedependent functions were derived from Andr6n and Paustian (1987), but, in contrast to their models, they include a temperature-independent coefficient (k~) for the litter fraction lost rapidly, and a temperature-dependent coefficient (k2) for the slowly decomposing fraction. Function 4 of Table 6 assumes a linear dependency on temperature, whereas the temperature effect in function 5 depends on a Q~o constant. To evaluate the effect of temperature, day-degrees were calculated as d

rsum= ~ T n=i

where T= temperature in degrees centigrade and d = day of exposure, or as d Qsum = E OIZO/10 n=l

For the function adjustment T~,m, and Qsum, were used, which are Ts.m/ TsummaxXyr and Q~um/Qs,,mmaxxyr, respectively, T~,,mrnaxand Q~mmax being the day-degrees at the end of the exposure and yr the years exposed. This does not change the form of the decomposition curve, but only converts the parameters to an annual basis instead of day-degrees. Water temperature (Fig. 3) was measured daily at the Biological Station of Illmitz at the eastern side ofthe lake. This station is at a considerable distance from the study area, but temperature only once differed by more than 1.5 °C from that measured at Site 3 when samples were collected, and usually by less than l ° C. All models were fitted to percentage AFDW remaining with the non-linear linefit procedure of Anonymous (1990). For the Q~,m model, Qto values of 1.5, 1.8, 2, 2.2, 2.5, 3, 4 and 5 were tested. In addition t o r E, r 2 was calculated as suggested by Kvhlseth (1985) to account for the number of model parameters (i)

NUTRIENTDYNAMICSOF PIIRAGMITES LITTER

223

TABLE 7 Results of the single exponential (k) and the T~ummodel (a, k~. k2) k

a

kl

k2

Ma~s remaining after I yeaP

0.757 1.125

12.57 21.19

>> 7.51

0.635 0.845

44.4 32.0

0.462 0.359 0.17 ! 0. ! 39

12.28 14.27 6.71 9.36

14.63 >> 20.24 6.75

0.346 0.226 0.118 0.068

60.7 67.4 g2.6 84.3

0.534 0.448

24.48 3 !.69

3.05 2. i I

0.313 0.185

55.3 59.9

0.153 0.102 0.119 0.094

27.62 26.59 4.46 3.71

0.55 -0.15 13.00 165.

0.050 0.170 0.084 0.065

84.6 92.1

0.569 0.617 0.556 0.397 0.485 0.395 0.812 0.890 0.903 0.472 0.499

1.85 5.45 5.73 12.21 13.85 12.83 34.08 20.27 21.23 0.002 8.70

>> 51.16 >> 5.90 7.52 9.52 2.75 4.38 3.70 0.29 1 !.56

0.533 0.533 0.472 0.244 0.305 0.230 0.397 0.620 0.631 0.452 0.377

57.8 52.8 56.3 67.3 61.7 67.8 44.9 40.7 40.0 61.0 60.4

0.341 0.142 0.135

14.53 5.18 4.78

8.38 >> 8.68

0.159 0.081 0.078

71.8 86.8 87.6

198 7 series

Site 1 Site 5 Site 3 Below water Leaves, coarse bag Leaves, fine bag Stalks, coarse bag Stalks, fine bag In reedless pool Leaves, coarse bag Leaves, fine bag Above water Leaves, coarse bag Leaves, fine bag Stalks, coarse bag Stalks, fine bag

89.9

1988 series

Leaves, coarse bags Site I --, 1 Site 3-,1 Site 5-, l Site 1--,3 Site 3--,3 Site 5--,3 Site ! ---,5 Site 3--,5 Site 5--,5 Site 2 Site 4 Site 3 Leaves, fine bag Stalks, coarse bag Stalks, fine bag

'Calculated from the results of the Tsummodel. >>, unreasonably large ( > 500).

r2 = 1- (1-r2)n/(n-i)

n being the number o f observations. The five functions o f Table 6 were adjusted to all data sets and average r z and r 2 are given below. As expected, the double exponential model resulted in better fits than the single exponential model. Average r 2 and r2 were 0.938

224

P. H|ETZ

1987 I-,. (.~

30

,.., 2 20-~ 15IO-

1q88

N DJ FMAM i i I I i I

5

1989

J JAS 0 N DJ FMAM i i I I i i I i i I I

-

~

1990

J JA5 I I I I

0 N DJ i I I I

F MA i I I

E 0N-5 . . . . . . . . . . .

,

. . . . . . . . . . .

i

. . . .

-100

-9o 18o '-7o

"""""

~ p o n . -ex o n e n t i o l ,

,.

E~

c

°"E ~ E}

E

. P . . .

I

r~ = 0 . 9 7 4 r= = 0.960 . . . .

,

~9~..~. . . . . . . . . . .

- _ ~D " 4 0 ,

....

'30

100-

90- ""

80-"

""

70-

a =,


6050-" 40-

--

30

"~--, T? ' ' m ' r= = 0 . 9 9 3

decoy coefficient , r' = 0.964 '

--

*

'

'

'

'

'

'

I

Osum, 0 1 0 = 2 , r ~ = 0.988

'

''---~.=..O '

'

'

'

'

'

,

,

,

I

'

'

~

models adjusted to AFDW

'

'

-lOO -9o -80 -7o -60 -so

'-40

-O~um, 01n=5, r~ = 0.997 . . . . . . . . . . . i . . . . . . . . . . . months exposed F i g . 3. R e s u l t s o f t h e d e c o m p o s i t i o n

,

,

. . . .

3 0

o f l e a v e s i n c o a r s e bags, e x p o s e d

at Site 3 in the reed stand below water; r 2 o f these adjustments are somewhat better than the average (see text). Daily water temperature was recorded at the Biological Station, IUmitz.

NUTRIENTDYNAMICSOFPHIUIGMITES LITTER

225

and 0.908, respectively, compared with 0.887 and 0.873, respectively, for the single exponential model. However, the resulting parameters were often difficult or impossible to interpret. Decay rates of the data sets differed considerably and sometimes even resulted in negative values. Equally, variable a, which can be interpreted as the labile fraction, was too variable to allow any identification with a chemical constituent. Mathematical procedures for nonlinear regression are generally evaluating least squares of the residuals and may converge to a local minimum, ignoring the absolute minimum and thus the best fit, which depends on the algorithm and starting value applied. This, therefore, supports the critique by De Lyon et al. (1983) and Moran et al. (1989), who suggested that multiple exponential models should not be used unless the various fractions have been identified. The decaying coefficient model, using only two constants, described the curves as well as the double exponential model (average r 2 and r 2 were 0.933 and 0.914, respectively). It assumes a decreasing decay rate, does not try to identify any material components and therefore does not suggest false interpretations. The average r 2 and r 2 for the Tsum model were 0.969 and 0.953, respectively. The Q~o values only had a little influence on the curve adjustment, but generally a Qmoof five gave the best fit (average r 2 and r 2 for the Qsem model were 0.969 and 0.954, respectively). Qlo values of biological processes are generally close to two (Andersen, 1978; Carpenter, 1982; Newell et al., 1985; Andr6n and Paustian, 1987). The Q~o calculated from soil respiration rates in the reed stand of Lake Neusiedl (Farahat and Nopp, 1966) was 2.1. The unusually high Q~o of five, its inconsistency with results from soil respiration data, and the fact that changing the Q~o had little influence on the adjustment, suggest that the Q~m model is not suitable for estimating the Qlo of the decomposition process, at least for the data presented here. The Tsum as well as the Qsum models tested appeared more suitable than the double exponential model to calculate the labile fraction. This was close to 10-20%o of dry weight for leaves and 6% fc - stalks and corresponds approximately to the AFDW lost during the first month of decomposition. It is likely that fraction a of the models in Table 6, which decreases quickly and independently of temperature, approximately equals the leachable components (Boyd, 1970), plus possibly part of the easily mineralised or depolymerised macromolecules. DISCUSSION

Assuming that the concentrations in green leaves sampled in August 1989 were similar to those present in August 1987 and 1988, changes in nutrient concentrations prior to leaf fall were calculated (Tables 4 and 5). Nitrogen, P, K and Mg decrease (in this order) relative to dry weight. These macronutrients can partially be retranslocated and stored (Sieghardt and Maier,

NUTRIENTDYNAMICSOF PHRAGMITES LITTER

227

macrofauna on reed litter decay in Lake Neusiedl has been evaluated by De la Cruz Mendez (1987). Temperature is an important factor controlling decomposition and can easily be measured and evaluated in aquatic environments where the influence of soil humidity need not be considered. However, oxygen concentration may have a significant effect (Gale and Gilmour, 1988) and can reverse tile enhancing effect of temperature, if increased respiration rates lead to an oxygen depletion limiting decomposition. In the case of the reedswamp studied, algal photosynthesis is over-compensating respiration and, at least in the water above the litter layer, oxygen coacentration is increasing rather than decreasing. Increased oxygen concentrations with temperature may explain why Qio values of the Qsum model are unusually high. Generally, decomposition was slow in the middle of the reed stand (Sites 2-4), higher near the landward margin (Site 1 ) and highest close to the open lake (Site 5). As decomposition was not influenced by differences in the initial mineral composition, the differences must be caused by physical, chemical, or biological characteristics of the sites. The shallow water at Site i (approximately 10 cm) and wave action stirring the deeper water at the lakeward margin (Site 5 ) may be causing a better oxygen supply for decomposers. Equally, oxygen supplied by wind action and photosynthesis from submerged macrophytes may explain the faster decomposition in the reedless pool at Site 3. Comminution due to wave action was probably not a major factor in the present study, as it is unlikely that the tough reed leaves confined in litterbags were damaged by being tossed around. Nitrogen is well known to limit decomposition processes (e.g. Kaushik and Hynes, 1971; Carpenter and Adams, 1979; Swift et al., 1979). Considering the low initial concentrations and the fact that N kept rising continuously, N limitation to microbial growth seems to prevail for several years. A possible influence of other nutrients on mineralisation is less well established. During decomposition, all nutrients increased considerably at the lakeward and landward margins of the reed stand and most nutrients remained comparatively low in the central part. Therefore, whatever nutrient may be critical, differences in their concentrations can account for at least part of the differences in the decomposition rates observed. As the translocation experiment showed, the effect does not depend on the initial nutrient concentrations, but rather on those which establish after a short time at the site of exposure. Frasco and Good ( 1982 ) exposed Spartina patens (Ait.) Muhl. litter in the Spartina alterniflora Loisel. zone, and vice versa, in a New Jersey salt marsh, and found weight losses depending on site as well as material. Obviously, in the present study, differences between litter from different sites, but from the same species, were not significant enough to result in different decay rates. The reed belt, at least in the area studied, is most eutrophicated in its landward and lakeward margins, because of the surrounding

226

P.HIETZ

1985; Smith et al., 1988). Calcium is important too, but is not mobile in the phloem. Iron and Na even seem to be accumulated in the senescent leaves, Fe because it would require much metabolic energy to retranslocate as Fe-chelates and Na because it is not needed in significant concentrations. Mineral element release in the first month of exposure can be attributed mainly to leaching (occurring in the first few hours to days) (Mason and Bryant, 1975; Gale and Gilmour, 1988). Release (or adsorption, in the case of Fe) of the cations seems to primarily reflect the mobility resulting from ion exchange processes (Davis and Van der Valk, 1978). Monovalent alkali cations are more mobile than bivalent earth alkali cations, and trivalent Fe is least mobile. Larger ions with the same charge are generally bound more strongly to negative groups. However, as the lake water contains more Na and Mg than K and Ca, Na and Mg accumulate relatively faster than the larger, less mobile ions. Subsequent cation accumulation, too, mainly reflects ion exchange processes and, to a minor extent, concentrations in the lake water. Uptake by litter-settling micro-organisms seems to be of little importance for cation accumulation, as the micro-organisms would incorporate nutrients according to their requirements (more K and Ca, less Na and Fe) rather than according to ion exchange patterns. At Site 5 the incorporation of sediment into the litter was of prime importance for the accumulation of Ca, Mg, and Fe. In leaves exposed in the reedless pool, more Ca and Mg accumulated than in the reed stand. In the pool, photosynthesis of submerged plants, mainly Characeae, removes CO2 from the water. This shifts the carbonate-bicarbonate equilibrium towards carbonate and increases pH, thus resulting in carbonate precipitation and its increase in the litter (Davis and Van der Valk, 1978 ). As cation concentrations in the litter keep rising for more than 2 years without reaching a constant equilibrium with the lake water, it may be supposed that the number of cation exchange site~ has increased. These are formed in the process of decomposition, which ultimately leads to humification and to the formation of humic acids. Unlike the cations, N and P tended to accumulate faster in the summer months than in winter. Changes in N and P content of decomposing litter are generally attributed to microbial activity (Kaushik and Hynes, 1971; Godshalk and Wetzel, 1978; Best et al., 1982). An effect of temperature on nutrient uptake may indicate whether nutrient accumulations are caused by abiotic or biotic processes, but is difficult to test statistically.

Factors controlling decomposition rate Different decomposition rates between coarse and fine mesh bags are well established and attributed to direct and indirect influence of the litter fauna (Mason and Bryant, 1975; Harrison, 1977; Polunin, 1982). The effect of the

228

P.HIETZ

farmland and sediments deposited from the open lake, respectively, whereas its central parts are comparatively nutrient-poor. Eutrophication o f the lake, therefore, is not only controlled by filtering nutrients as the external influx passes through the wetland, but nutriems are also taken up as the sediment load o f the turbid water is deposited in the reed stand. Local differences in nutrient absorption should receive m o r e attention when estimating the potential o f wetlands for eutrophication control. During the past 200 years, as the reed stand has been attaining its present size, vast a m o u n t s o f nutrients have been incorporated into its litter layer o f several decimeters, preventing over-eutrophication. If the reed stand stopped extending, which m a y already have been caused by raising the water level in 1965, its nutrient-storage capacity will come to an end sooner or later. To stop the reed from extending may be desirable, as otherwise it may come to cover the whole lake. However, unless a feasible m a n a g e m e n t for a substantial nutrient export is found, it will be necessary to curb nutrient input drastically, or else the lake will face increasing eutrophication. This would still be enhanced if the reed was to decline as in other European lakes. ACKNOWLEDGEMENTS The author wishes to thank Professor R. Maier for his continuous encouragement and valuable c o m m e n t s throughout this study and to U. Roy-Seifert for corrections o f the English manuscript.

REFERENCES AktionsgemeinschafiGesamtkonzept NeusiedlerSee, 1985. Forschungsbericht1981-1984. Wiss. Arb. Burgenland, 72. Andersen, F.O., 1978. Effects of nutrient level on the decomposition of Phragmites communis Trin. Arch. Hydrobiol., 84: 42-54. Andr6n, O. and Paustian, K., 1987. Barley straw decomposition in the fieldua comparison of models. Ecology,68:1190-1200. Anonymous, 1990. SigmaPlot. Handbook. Jandel Scientific, Cone Madera, CA, USA. Best, E.P.H., Zippin, M. and Dassen, J.H.A., 1982. Studies on decomposition of Phragmites australis under laboratc::y,conditions. Hydrobiol. Bull., 16:21-33. Boyd, C.E., 1970. Lo~sesof mineral nutrients during decomposition of Typha latifolia. Arch. Hydrobiol., 65:511-517. Carpenter, S.R., 1982. Comparisons of equations for decay of leaf litter in tree-hole ecosystems. Oikos, 39: 17-22. Carpenter, S.R. and Adams, M.S., 1979. Effectsof nutrients and temperature on decomposition of Myriophyllum spicatum L. in a hard-water eutrophic lake. Limnol. Oceanogr., 24: 520528. Crossley, D.A. and Hoglund, M.P., 1962. A litter-bag method for the study of microanhropods inhabiting leaf litter. Ecology,43:571-574.

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Davis, C.B. and van der Valk, A.G., 1978. The decomposition of standing and fallen litter of Typl~a glauca and Scirpusfluviatiiis. Can. J. Bot., 56: 662-675. De la Cruz Mendez, M., 1987. Die Rolle der Makrofauna bei den Abbauprozessen yon Phragmites australis im Schilfgiirtel des Neusiedlersees. Ph.D. Thesis, University of Vienna., 122 pp. De Lyon, M.H., van Laar, E.M.J.M. and Brock, T.C.M., 1983. A comparison of three mathematical models for the description of breakdown of Nuphar lutea detritus. Proc. Int. Syrup. Aquatic Macrophytes, Nijmegen, Netherlands, 18-23 September 1983, pp. | 29-133. Farahat, h,.Z. and Nopp, H., 1966. Ober die Bodenatmung im Schilfgiirtel des Neusiedler Sees. Sitzurgsber. Oesterr. Akad. Wiss. Math. Naturwiss. Kl. Abt. I, 175: 237-255. Frasco, F;.A. and Good, R,E., 1982. Decomposition dynamics ofSpartina alterniflora and Spartinf. patens in a New Jersey salt marsh. Am. J. Bot., 69: 402-406. Gal~, P.,~,~.~,nd Gilmour, J.T., 1988. Nct mir,c~alizati m of carbon and nitrogen under anaerobic and aerobic conditions. Soil Sci. Soc. Am. J., 52:1006-1010. Godshalk, G.L. and Wetzel, R.G., 1978. Decomposition of aquatic angiosperms. If. Particulate components. Aquat. Bot., 5: 301-327. Gunatilaka, A., 1985. Niihrstoffkreisl~iufe im Sehilfgiirtel des Neusiedler Sees. Auswirkungen des Grtinschnittes. Wiss. Arb. Bgld., 72:223-310. Harrison, P.G., 1977. Decomposition of macrophyte detritus in seawater: effects of grazing by amphipods. Oikos, 28:165-169. Hietz, P., 1991. Freisetzung yon N~ihrstoffen dutch Abbau von Phragmites australis im Schilfgiirtel des Neusiedlersees. D.Sc. Thesis, University of Vienna, 169 pp. lmhof, G. and Burian, K., 1972. Energy-flow Studies in a Wetland Ecosystem. Special Publication ofthe Austrian Academy of Sciences for the IBP. Springer, New York, pp. 2-15. Kaushik, N.K. and Hynes, H.B.N., 1971. The fate of the dead leaves that fall into streams. Arch. Hydrobiol., 68:465-515. Kv~tlseth, T.O., 1985. Cautionary note about R 2. Am. Stat., 39: 279-285. L6ffler, H. (Editor), 1979. Neusiedlersee: The Limnology of a Shallow Lake in Central Europe. Junk, The Hague, 534 pp. Mason, C.F. and Bryant, R.J., 1975. Production, nutrient content and decomposition of Phragmites communis Trin. and Typha angustifolia L. J. Ecol., 63: 71-95. Moran, M.A., Benner, R. and Hodson, R.E., 1989. Kinetics of microbial degradation of vascular plant material in two wetland ecosystems. Oecologia, 79: 158-167. Neuhuber, E, Brossmann, H. and Zahradnik, P., 1979. Phosphorus and nitrogen. In: H. L6ffier (Editor), Neusiedlersee: The Lumnology of a Shallow Lake in Central Europe. Junk, The Hague, 534 pp. Neweil, S.Y., Fallon, R.D., Calrodriguez, R.M. and Groene, L.C., 1985. Influence of rain, tidal wetting and relative humidity on release of carbon dioxide by standing-dead salt-marsh plants. Oecologia, 68: 73-79. Olson, J.S., 1963. Energy storage and the balance of producers and decomposers in ecological systems. Ecology, 44:322-331. Ostendorp, W., 1989. 'Die-back' of reeds in Europe - a critical review of literature. Aquat. Bot. 35: 5-26. Polunin, N.V.C., 1982. Effects of the freshwater gastropod Planorbis carinatus on reed (Phragmites australis) litter microbial activity in an experimental system. Freshwater Biol., 12: 547-552. Reddy, K.R. and Smith, W.H. (Editors), 1987. Aquatic Plants for Water Treatment and Resource Recovery. Magnolia, Publ., Orlando, FL, 1032 pp. Sachs, L., 1978. Angewandte Statistik. 5. Aufl. Springer, Berlin, 523 pp. Sieghardt, H. and Maier, R., 1985. Produktionsbiologische Untersuchungen an PhragmitesBesfiinden im geschlossenen Sehilfgih'tel des Neusiedler Sees. Wiss. Arb. Bgld., 72:189-221. Smith, C.S., Adams, M.S. and Gustafson, T.D., 1988. The importance ofbelowground mineral element stores in cattails (Typha latifolia L.). Aquat. Bot., 30: 343-352.

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Statistical Analysis Systems Institute, 1985. SAS User's Guide, Statistics. Version 5. SAS Institute Inc., Cary, NC. Swift, M.J., Heal, O.W. and Anderson, J.M., 1979. Decomposition in Terrestrial Ecosystems. Studies in Ecology, Vol. 5. BlackweU Scientific, Oxford, 431 pp. Van der Emde, W., Matsch6, N. and Plahl-Wabnegg, F., 1985. EinfluB yon Hochwasserereignissen auf die N~ihrstoffbelastung der Wulka und deren Auswirkungen auf die Stoffumsetzungen im Schilfgiirtel des Neusiedler Sees. Wiss. Arb. Bgld., 72:9 l-121. Wieder, R.K. and Lang, G.E., 1982. A critique of the analytical methods used in examining decomposition data obtained from litter bags. Ecology, 63:1636-1642. Zar. J.H., 1984. Biostatistical Analysis. 2rid edn. Prentice-Hall, Englewood Cliffs, N J, 714 pp.