Controls on soil cellulose decomposition along a salinitygradient in aPhragmites australiswetland in Denmark
Irving A. Mendelssohna,∗, Brian K. Sorrellb,1,Hans Brixb,Hans-Henrik Schierupb, Bent Lorenzenb, Edward Maltbyc
a Wetland Biogeochemistry Institute, Louisiana State University, Baton Rouge, LA 70803, USAb Department of Plant Ecology, Aarhus University, Nordlandsvej 68, Risskov DK-8240, Denmark
c Royal Hollaway College, University of London, Huntersdale Callow Hill, Virginia Water,Surrey GU25 4LN, UK
Abstract
Although soil organic matter decomposition is an important process determining nutrient trans-formations and availability in wetland ecosystems, few studies have attempted to assess whichenvironmental factors are most important in controlling spatial differences in decomposition ratesfound along environmental gradients. Relative soil decomposition was determined in aPhragmitesaustralisCav. Trin ex Steudel dominated wetland in northern Jutland, Denmark along a naturalsalinity gradient, where nutrients, soil moisture, temperature and salinity among other factors alsovaried. Our objective was to identify which edaphic factors most limited rates of relative soil de-composition, as evaluated by measuring cellulose decomposition with the cotton strip technique.Replicate cotton strips were placed at seven marsh sites along the salinity gradient, and soil and in-terstitial water samples were collected and analyzed for major macro- and micronutrients (NH4–N,NO3–N, P, PO4, K, Mg, Ca, Na, S, Fe, Mn, Zn, Cu, Mo, B, Si), pH, Eh, conductivity, temperature,and soluble sulfides. Cellulose decomposition, expressed as cotton tensile strength loss (CTSL)per day, decreased with increasing salinity, except at the highest salinity site where a significantincrease occurred. Mean CTSL values, averaged for each marsh site, varied 3-fold from 1.8 to5.5% loss per day. Principal component and multiple regression analyses were used to prioritizethe importance of the various factors that might control this spatial difference in CTSL rates. Al-though soil conductivity (salinity) accounted for the large percentage (45%) of the variation in theenvironmental data, soil fertility- and soil reduction-associated variables explained the greatest per-centage (56%) of the spatial variation in cellulose decomposition. Univariate correlation analysessupported the conclusion that soil fertility, primarily inorganic nitrogen and phosphorus, is the major
Wetlands, which worldwide comprise 7–8 million km2 (Mitsch et al., 1994), are nowrecognized as highly valued ecosystems (Costanza et al., 1998) that provide important soci-etal services such as pollution abatement, flood control, shoreline protection, groundwaterrecharge and others (Mitsch and Gosselink, 1993). These systems, however, are being per-turbed by a myriad of anthropogenic activities including hydrologic modification, dredgeand fill activities, grazing and harvesting, nutrient runoff, and pollutant input (Williams,1993; Viles and Spencer, 1995). In addition, the predicted effects of global warming in ac-celerating sea level rise are likely to further impact coastal wetlands by promoting salinityintrusion and submergence (Titus, 1988).
A number of ecological processes could be affected by these environmental changes. Thedecomposition of soil organic matter, which controls the mobilization of nutrients for plantgrowth, the transformations of nutrients and pollutants important in pollution abatement,and the turnover and accretion of organic matter enabling marshes to vertically accrete andkeep pace with water level rise, may have some of the most far-reaching implications rela-tive to environmentally induced effects on wetland stability and sustainability. Although thedecomposition of organic matter in wetlands has been extensively examined, few studieshave specifically addressed the factors that control site-specific differences in rates of de-composition, especially from a multi-factor perspective (for reviews see Good et al., 1982and Webster and Benfield, 1986).
A number of abiotic factors may influence organic matter decomposition by microbiotain the wetland environment including (1) nutrients, (2) oxygen, (3) hydroperiod, (4) temper-ature, (5) salinity, and (6) pH. Biotic factors such as plant litter quality and faunal activityare also important. Although nutrients and oxygen are likely to be the major abiotic con-trollers of site-differences in organic matter decomposition within wetlands (temperaturebeing important in controlling seasonal and annual variability), surprisingly few studieshave attempted to empirically differentiate which of these factors actually determine or-ganic matter decomposition in the field (Latter and Harrison, 1988). The need to identifythe factors controlling wetland organic matter decomposition has recently become impor-tant in the context of the extensive European die-back of the common reed,Phragmitesaustralis(Ostendorp, 1989; Cizkova-Koncalova et al., 1992; Ostendorp et al., 1995). Eu-trophication and water level increases may be partially responsible for this die-back byaltering soil organic matter decomposition and consequently generating toxic compoundssuch as organic acids and sulfides (Armstrong et al., 1996a, b). CoastalPhragmitesmarshes,which will be subject to elevated salinities as well as higher water levels from acceleratedsea level rise, may be particularly vulnerable. However, the effects of environmental factorssuch as salinity, soil aeration, and nutrients in modifying soil organic matter decompositionin Phragmiteswetlands have not been investigated simultaneously in an attempt to explain
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Fig. 1. Map of the eastern section of the Vejlerne Nature Reserve with sampling locations; inset map shows relationto northern Jutland, Denmark.
site-specific differences in soil decomposition. Thus, the objective of this study was to elu-cidate the relationship between soil organic matter decomposition rate, as determined bythe cotton strip technique (Latter and Howson, 1977), and soil environmental conditionsalong a salinity gradient in a coastalPhragmiteswetland in Northern Jutland, Denmark.
2. Materials and methods
2.1. Study site
Field work was carried out during the summer of 1994 in the Vejlerne Nature Reserve,a 150 km2 coastal wetland adjacent to the Limfjord in northern Jutland, Denmark (Fig. 1).The area originally consisted of two shallow saltwater fjords, which were separated from theLimfjord during the 1870s by the construction of dikes in an unsuccessful attempt to reclaimland for agriculture. Subsequent infiltration of fresh groundwater has resulted in a mosaic ofdiffering salinities and a proliferation of wetland macrophytes, especially extensive standsof P. australis. The sediment varies from a eutrophic organic mud to a mixture of mud andsand overlying the original nutrient-poor marine sand.
We studied decomposition at seven marsh sites colonized by reed in Han Vejle andBygholm Vejle, near the eastern boundary of the wetland (Fig. 1). All sites were accessiblefrom a nearby causeway adjacent to the wetland and were chosen to provide a range ofsalinity. Sites 1–3 were located in an oligohaline reed swamp east of the causeway in HanVejle. Site 1 was an open lagoon in which the reeds had died, leaving a ca. 0.50 m-deep
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sediment still containing old rhizome material from earlier reed growth. Site 2, immediatelyadjacent, had a similar sediment but still contained some live reed, whereas Site 3 was anearby vigorous, healthy reed stand unaffected by die-back. Sites 4–7, west of the causewayin Bygholm Vejle, followed a gradient of increasing salinity towards the Limfjord.
2.2. Sampling design
At each of the seven reed stand sites, seven individual sampling locations were randomlyselected along transects 10 to 30 m in length depending on the size of thePhragmitesstand.At each of the sampling locations, cellulose decomposition and selected environmentalvariables were measured as described below.
2.3. Decomposition rate
Decomposition of cellulose in cotton strips was used as a proxy for evaluating the rel-ative rates of decomposition of organic matter in the sevenPhragmitesstands (Latter andHowson, 1977; Harrison et al., 1988). Cellulose comprises about 70% of the organic carboncompounds in plant tissue and its decay rate is, therefore, a key factor in plant decompo-sition. The cotton strip technique for quantifying cellulose decomposition is based on theloss of tensile strength of cellulose fibers, referred to as cotton tensile strength loss (CTSL),of a standardized cotton fabric comprising 97% holocellulose. This technique has been ex-tensively used in a variety of environments, including wetlands, as a quantitative measureof cellulose decomposition and a relative measure of soil decomposition (see French, 1988and Harrison et al., 1988).
At each of the seven sampling locations within each of the seven reed stands, a strip of thestandard cotton material (Shirley Institute, Didsbury, Manchester, UK) of 12 cm× 30 cmwas inserted vertically into the soil substrate with the aid of a sharpshooter shovel asdescribed by Maltby (1988). Cotton strips of 12 cm× 90 cm were also employed at marshSites 1 through 3 to assess cellulose decomposition at greater depths. The soil surface levelwas marked on each strip with a small lateral cut (ca. 6 cm of the cotton strip was above thesoil surface). The strips were placed in the marshes on 21 June, 1994 and retrieved on 6 July,1994 after a 15-day exposure period. Reference cotton strips, used to quantify the tensilestrength of the non-decomposed material, were inserted in the soil and immediately retrievedon 6 July, 1994. After retrieval, all strips were immediately washed in freshwater from anadjacent pond to remove the majority of soil and debris and washed again in deionizedwater. The samples were dried at 33◦C and stored in plastic bags until analysis. In thelaboratory, the strips were cut into horizontal segments 3 cm wide and reduced by frayingto 2 cm segments that resulted in test units corresponding to depths of 0–2, 3–5, 6–8 cm, etc.Tensile strength (TS) was measured with a tensometer (Monsanto Type W) equipped with7.5 cm wide jaws adjusted to 3 cm spacing. All measurements were carried out at 18–22◦Cand 100% humidity obtained by soaking the strips in deionized water. Individual and meanloss in TS were calculated relative to the mean of the reference cotton strips. These datawere used to calculate cotton tensile strength loss (CTSL) on a percent loss per day basisfor each level in the profile. Losses in tensile strength were averaged over a soil depth of0 to 22 cm to provide an integrated value of loss in tensile strength that could be related tothe soil environmental variables.
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2.4. Soil characterization
At each of the seven marsh sites, five samples of the soil substrate were taken with a 10 cmdiameter stainless steel corer (35 soil samples). The corer was cut into the substrate to a depthof 40–60 cm, filled with water, plugged and then withdrawn. The soil sample was pressedout of the corer using a piston, and the 10–20 cm depth fraction was carefully transferredto a 1 l plastic bottle with lid. Care was taken not to lose any material – including water –during this process. In the laboratory, the samples were weighed (fresh weight), dried to aconstant weight at 105◦C and again weighed for the determination of water content. Next,the samples were homogenized, and two subsamples ignited at 550◦C for determinationof loss on ignition (LOI) as a measure of organic matter content. Sub-samples (0.5 g) ofthe remaining ash material were extracted in boiling 1 mol l−1 HCl for 10 min, and the soilP concentrations analyzed on ICP (Plasma II Emission Spectrometer, Perkin Elmer) afteradequate dilution in 0.1 mol l−1 HCl. Total soil nitrogen was analyzed as Kjeldahl-N usingstandard procedures.
2.5. Interstitial water analyses
A suction sampler was used to remove interstitial water from the 10–20 cm zone of thesubstrate as described previously (McKee et al., 1988; Koch and Mendelssohn, 1989). Thesampler consisted of a small diameter (3 mm inside diameter) rigid plastic tube, containingnumerous ca. 0.5 mm diameter holes covered with 3 to 4 layers of cheesecloth, connectedvia tygon tubing and a 3-way valve to a 50 ml syringe. This apparatus allowed the collec-tion of ca. 50 ml of relatively clear interstitial water from the soil within a few secondswithout significant exposure to the atmosphere. A 5 ml unfiltered aliquot was immediatelytransferred through the three-way valve to an equal volume of antioxidant buffer (Lazar op-erating instructions for Model IS-146 Sulfide Electrode) for measurement of total solublesulfide with a sulfide electrode (Lazar Research Laboratories, Los Angeles, CA). A secondunfiltered aliquot was collected for pH (Cole-Parmer Digi-Sense portable pH meter andelectrode) and conductivity (Knick conductivity meter 702, Berlin, Germany) determina-tions. The final aliquot was filtered through a 0.45m uniflow in-line filter and acidified witheither nitric acid for the analysis of major elements (P, K, Ca, Mg, Na, S) and trace metals(Fe, Mn, Zn, Cu, B, Si, Mo) (ICP-AES, Perkin Elmer Plasma 2000) or hydrochloric acidfor ammonium, nitrate, and phosphate determinations (Lachat Quikchem 4-channel flowinjection analyzer, Milwaukee, WI).
2.6. Soil redox potential and temperature
Soil redox potentials (Eh) were measured 15 cm below the soil surface with bright plat-inum electrodes that were allowed to equilibrate for approximately 30 min prior to measure-ment. Each electrode was checked before use with quinhydrone in pH 4 and 7 buffers (mVreading for quinhydrone is 218 and 40.8, respectively, at 25◦C). The potential of a calomelreference electrode (+244 mV) was added to each value to calculate Eh. Temperature wasmeasured at 15 cm below the soil surface with a digital thermometer.
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Table 1Selected soil characteristics at the sevenPhragmites australis-dominated study sites located in the Vejlerne NatureReserve, Jutland, Denmarka
Marsh sites Soil variables
Soil moisture Bulk density Organic matter Total N Total P(% wet wt ) (g cm−3) (% dry wt) (mmol g−1 dry wt) (mmol g−1 dry wt)
1 91.4± 1.9 a 1.05± 0.02 a 29.4± 4.0 a 1126± 182 a 56± 10 a2 92.2± 0.5 a 1.11± 0.09 a 60.9± 2.4 b 1728± 93 b 132± 11 b3 92.5± 0.2 a 0.87± 0.10 b 74.5± 1.6 c 1751± 49 b 169± 19 b4 91.6± 0.7 a 0.99± 0.03 ab 66.1± 7.5 bc 1647± 89 b 165± 43 b5 79.2± 0.7 b 0.80± 0.02 bc 25.9± 2.7 a 628± 98 c 26± 5 a6 50.9± 1.3 c 1.34± 0.04 d 8.0± 0.7 d 189± 28 d 10± 1 a7 44.5± 2.8 d 1.53± 0.10 d 2.7± 0.4 d 82± 12 d 5± 0.3 a
a Values are means± 1 SE (n= 5) and means with different letters are significantly different atP= 0.05.
2.7. Statistical analyses
The environmental data (interstitial variables and soil Eh and temperature), which weremeasured to identify differences amongPhragmitesstands and their influence on decom-position rate, were first collectively analyzed with a Multivariate Analysis of Variance(MANOVA). After a significant multivariate treatment affect was detected by the WilksLambda test, individual univariate tests were performed on all abiotic data. The univari-ate model was a completely randomized one-way analysis of variance. The Tukey’s HSD(honestly significant difference) test was used to identify significant differences amongtreatments. Unless otherwise noted, all significant differences are at a probability of 0.05or less.
A rotated principal components analysis was conducted on the environmental data andthe resulting factor scores used as the independent variables in a multiple regression withCTSL (% loss per day) as the dependent variable. This procedure provided the ability toreduce the number of independent variables from 20 individual abiotic variables to threeprincipal component factors that are independent linear combinations of the 20 originalvariables.
All data were analyzed with JMP for the Macintosh (SAS Institute Inc., 1994).
3. Results
3.1. Site characterization
The values of the soil variables were different among the sevenPhragmites-dominatedmarsh sites (Table 1). Soil moisture was significantly greater at Sites 1 through 4 comparedto Sites 5 through 7, which differed significantly from each other. Similarly, LOI was highestat Sites 2, 3 and 4 compared to the remaining sites. Soil bulk density was relatively highat all sites with mean values at least 0.8 g cm−3. However, bulk density exhibited a markedincrease at Sites 6 and 7 compared to the rest of the marsh study sites. Total N and total Pmirrored the trends in LOI values and varied from relatively high levels at Sites 1 through4 to significantly lower levels at Sites 5 through 7.
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3.2. Decomposition rates
3.2.1. Site VariationCellulose decomposition differed significantly along the salinity gradient. CTSL, aver-
aged over a depth of 22 cm for each cotton strip, was highest at Sites 1 and 2, progressivelylower through to Site 6, and significantly higher at Site 7 compared to Sites 5 and 6 butsignificantly lower than for Sites 1 and 2 (Fig. 2(A)). CTSL ranged from a high of 5.5% perday at Site 2 to a low of 1.8% per day at Site 6. The differences in CTSL among sites wereevident throughout much of the depth profile (Fig. 3(A)). Marsh Sites 5 and 6 has some ofthe lowest soil profile decomposition rates and marsh Sites 1 and 2 some of the highest,while Sites 3, 4 and 7 were intermediate.
3.2.2. Depth variationCellulose decomposition rates decreased with depth, as exemplified at marsh Sites 1, 2
and 3, where CTSL rates were determined to a depth of 74 cm (Fig. 3(B)). CTSL rates werehigher than 5% per day near the soil surface, and decreased significantly to as low as ca. 1%per day at the bottom of the profile. Decomposition rates decreased with depth even whenthe maximum depth was a more shallow 22 cm and differences in site decomposition weregreat (Fig. 3(A)). Although marsh Sites 1 and 2 showed no significant decrease in cellulosedecomposition to a depth of 22 cm, CTSL at these sites significantly decreased at depthsgreater than 22 cm (Fig. 3(B)). At marsh Sites 4, 5 and 6, there existed a tendency, albeitnon-significant, for higher rates of cellulose decomposition at depths of 4 to 7 cm comparedto depths above or below in the profile (Fig. 3(A)).
3.3. Environmental data
A Multiple Analysis of Variance (MANOVA) of the 20 environmental variables showeda highly significant (P< 0.0001) treatment effect based on the Wilks’ Lambda test andtherefore, the environmental data, as a whole, significantly differed among sites. As a result,ANOVA was conducted on each of the individual variables and the results discussed below.
3.3.1. NutrientsInterstitial NH4 + NO3, P, K, Ca, Mg, Na and S significantly varied with marsh site along
the salinity gradient (Fig. 2(B–H)). Nitrogen concentration (NH4 + NO3) was highest at Site1, significantly lower at Site 2, and lowest at the remaining marsh sites (3 through 7) whichdid not significantly differ from each other (Fig. 2(B)). Phosphorus concentration varied ina similar fashion to N, with the exception that the P concentration was significantly higher atSite 7 compared to the low P levels at Sites 4–6 (Fig. 2(C)). Potassium, Mg, Na and S (Fig.2(D, E, G, and H) respectively) and Mo (Fig. 4(C)) varied similarly over the salinity gradientwith the lowest concentrations in the freshwater Sites 1 through 4, and greater values fromSites 5 through 7 at the higher end of the salinity gradient. The micronutrients Fe and Mnexhibited similar trends along the salinity gradient with lowest concentrations in the freshersegment of the gradient and higher concentrations in the more saline portion (Fig. 4(A andB)). However, within-site variation was high, especially for iron. Nonetheless, significantdifferences for both Fe and Mn occurred among the study sites with Fe and Mn significantly
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Fig. 2. Cellulose decomposition (0–22 cm soil depth) (A) and interstitial water concentrations of ni-trate + ammonium (B), phosphorus (C), potassium (D), magnesium (E), calcium (F), sodium (G), and sulfur(H) at the sevenPhragmitesmarsh sites located along a salinity gradient in northern Denmark (Mean± 1 SE,n= 7). Means with different letters are significantly different atP= 0.05.
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Fig. 3. Cellulose decomposition with soil depth (22 cm) at the sevenPhragmitesmarsh sites (A) and to a soil depthof 75 cm at the threePhragmitesmarsh sites with the lowest salinities (B) along a salinity gradient in northernDenmark (Mean± 1 SE,n= 7).
higher at Site 7 compared to Sites 1 through 6. Although Zn concentrations were generallyhighest at Sites 6 and 7, lowest at Sites 3, 4 and 5, and intermediate at Sites 1 and 2 (Fig.4(E)), these differences were not statistically different. Copper concentrations were belowdetection limits. Silicon concentration was progressively lower from Site 1 to Site 5 andthen significantly higher at Sites 6 and 7 (Fig. 4(F)). Boron concentration, however, wasprogressively lower from Site 1 to Site 3, significantly higher at Sites 4 and 5, still higherat Site 6, but significantly lower at Site 7 (Fig. 4(D)).
3.3.2. Eh, Sulfide and pHSoil redox potential generally increased along the salinity gradient (Fig. 5(A)). Eh was
lowest at Site 1, where die-back had occurred and no living vegetation was present, signif-icantly higher at Sites 2, 3, and 4, still higher at Site 5, lower at Site 6 to a level similar to
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Fig. 4. Interstitial water concentrations of iron (A), manganese (B), molybdenum (C), boron (D), zinc (E), andsilicon (F) at the sevenPhragmitesmarsh sites located along a salinity gradient in northern Denmark (Mean± 1SE,n=7). Means with different letters are significantly different atP= 0.05.
that of Site 4, and again significantly higher at Site 7. Sulfide showed a different trend withhighest concentrations at Site 1 and significantly lower levels through Site 4 and thereafterremaining statistically similar at Sites 4 through 7 (Fig. 5(B)). Sites 1–3 had the highestpH’s, with Site 4 intermediate, and Site 5–7 lowest (Fig. 5(C)). On average, interstitial pHvaried ca. 1 unit along the salinity gradient.
3.3.3. Conductivity and temperatureConductivity changed as expected along the salinity gradient (Fig. 5(D)). Sites 1–3 had
the lowest conductivities which did not significantly differ among themselves. Conductivitywas significantly higher at Site 5 than Sites 1 through 4 and still higher at Site 7, wheresalinities reached ca. 43% of full strength seawater (35 ppt = 50 mS cm−1 at 20◦C). Mean
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Fig. 5. Soil redox potential (A), interstitial water sulfide (B), pH (C), conductivity (D) and temperature (E) at thesevenPhragmitesmarsh sites located along a salinity gradient in northern Denmark (Mean± 1 SE,n= 7). Meanswith different letters are significantly different atP= 0.05.
temperature varied from 13 to 16◦C among sites (Fig. 5(E)). Although some significantdifferences among sites occurred, this relatively small variation in temperature had norelationship to cellulose decomposition (r =−0.01,P> 0.05).
3.4. Decomposition and environmental controls
The 20 environmental variables that were collected at seven sampling locations withineach of the seven marsh sites were simplified into three major principal components de-scribing the variation in the environmental data (Table 2). The first principal componentexplained the largest percentage of the variation (45%) and had high loadings for conduc-tivity, K, Na, and Mg (Table 2) and can be interpreted as a salinity-related factor. Principalcomponent 2 explained an additional 25% of the variation and had high loadings for NH4,P, and sulfide (Table 2); this factor is interpretable as a nutrient-soil reduction related factor.Principal component 3 explained an additional 10% of the variation and had high loadings
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Table 2Principal components analysis of the 20 environmental variables (interstitial nutrients and sulfide, soil conductivityand temperature) at the 49 sampling locations (7 marsh sites× 7 replicates per site) within the Vejlerne NatureReserve, Jutland, Denmarka
Eigenvalue Proportion of variance (%) Cumulative proportion of variance (%)
a Variables with high loadings are shown in boldface type.
for B, S, and Ca; this factor is also likely related to seawater given the significant positivecorrelations between these elements and conductivity (r = 0.53,r = 0.76, andr = 0.71, re-spectively;P< 0.01). The three interpretable factors from the principal component analysisexplained 79% of the variation in the environmental data with salinity and soil nutrients-soilreduction explaining the greatest amount of the variation.
To identify which environmental factors were most related to differences in cellulose de-composition along this salinity gradient, the three principal components that were identifiedwith the principal component analysis and explained the greatest percentage in the variationof the environmental data, were used as independent variables in a multiple regression, wherethe dependent variable was CTSL. Two of the three principal components were significantin the multiple regression. Principal component 2, the nutrient-soil reduction factor, enteredinto the multiple regression first and accounted for 56% of the variation in % loss of tensilestrength. Principal component 3, the secondary salinity factor comprising boron, calciumand sulfur, accounted for an additional 21% of the variation in cellulose decomposition.
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Thus, a total of 77% of the variation in cellulose decomposition could be explained by thesefactors. A univariate correlation analysis of the environmental data with CTSL showed thatPO4 had the highest correlation with CTSL (r = 0.76,P< 0.01); NH4 + NO3 and sulfide hadthe next highest correlations (r = 0.65,P< 0.01;r = 0.68,P< 0.01, respectively).
4. Discussion
Decomposition of plant organic matter is a complex process involving both intrinsic andextrinsic controls (Webster and Benfield, 1986; Colberg, 1988). Intrinsic controls include thequality, i.e., chemical composition, of the decomposing organic material, while extrinsiccontrols encompass faunal and microbial effects as well as the influence of the abioticenvironment on these biotic populations. Soil fertility, moisture, temperature, oxygen, pHare some of the important extrinsic abiotic variables that affect decomposition rate.
This research utilized a standard cellulose material by which to assess soil organic matterdecomposition in aPhragmiteswetland. Because the cotton strips are virtually identical,any difference in decomposition rate will be due to extrinsic variables. Mean cellulosedecomposition, measured as loss of tensile strength of the Shirley Institute cotton material,varied 3-fold over the seven marsh sites sampled along the salinity gradient while individualmean CTSL values varied as much as 10-fold. What abiotic factors produced such a largevariation in decomposition rate?
Although soil interstitial water conductivity varied from near fresh water to approxi-mately half sea-strength, interstitial conductivity was not significantly correlated with cel-lulose decomposition (r =−0.20P> 0.05). Also, the conductivity-related variables (Princi-pal Component 1) from the principal component analysis did not account for a significantpercentage of the spatial variation in cellulose decomposition in the multiple regression.However, Principal Component 3, which explained 21% of the spatial variation in cellu-lose decomposition, contained some of the elements which are found in relatively highconcentrations in seawater such as boron, calcium and sulfur (note that interstitial S waspositively correlated with conductivity [r = 0.76,P< 0.01] and negatively correlated withinterstitial sulfide [r =−0.35,P< 0.05], suggesting that the interstitial S was primarily sul-fate). These elements exhibit significant negative correlations with cellulose decompositionrate (B =−0.56,P< 0.01; Ca =−0.35,P< 0.05; S =−0.50,P< 0.01). Thus, it appears thatseawater may have contributed to the spatial variation in decomposition, but the mecha-nism(s) for this response was not likely due to the dominant seawater cations Na, K, and Mgnor their integrator, conductivity, because these variables were not significantly correlatedwith cellulose decomposition (r =−0.15, r = 0.05,r =−0.20 andr =−0.20, respectively;P> 0.05) nor was Principal Component 1, on which these variables were highly loaded,significantly related to cellulose decomposition. Rather, seawater appears to negatively in-fluence cellulose decomposition through boron, sulfur, calcium or other factors correlatedwith these elements. Calcium, H3BO3 and SO4
2− are major constituents of seawater, andwe hypothesize that, within the conductivity range of this investigation, some of these orassociated variables may be exerting their influence on decomposition. Further researchis required to test this hypothesis. The negative effect of salinity on the decomposition ofPhragmitesleaves has been indicated in three stream ecosystems that differed primarily insalinity (from 2 to 13 ppt) (Reice and Herbst, 1982). Here, high salinity was thought to have
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inhibited the growth of bacteria and fungi on the detritus. However, Tanaka and Tezuka(1982) demonstrated that both free-living and attached bacteria can colonizePhragmitesorganic matter when cultured in seawater. Furthermore, the decomposition of mangrove(Bruguiera gymnorrhiza) leaves, although showing a significant reduction in decomposi-tion rate at 30 ppt compared to 0 ppt, was not significantly different between 15 and 0 pptwhen grown under controlled laboratory conditions (Steinke and Charles, 1986). Thus, itmay not be surprising that a conductivity gradient of approximately 2 to 23 mS cm−1 (1to 17 ppt) in the present study had no direct effect on cellulose decomposition. However,an indirect effect of salinity via the variables associated with Principal Component 3 can-not be discounted. Although temperature differences can have dramatic effects on organicmatter decomposition (Godshalk and Wetzel, 1978; Carpenter and Adams, 1979; Nowicki,1994; Updegraff et al., 1995), the differences in soil temperature among the marsh siteswere not large enough to contribute to the observed spatial variation in decomposition rateas evidenced by the absence of a significant correlation between temperature and cellu-lose decomposition (r =−0.01,P> 0.05). Similarly, Neckles and Neill (1994) found thattemperature did not explain spatial variation in decomposition rates in a fresh marsh inCanada.
Nutrient concentration is a primary factor influencing organic matter decomposition.Furthermore, oxygen concentration and/or oxidation–reduction intensity (Eh), especiallyin wetland systems, may exert an additional control over decomposition rates. Both soilfertility- and redox-related variables accounted for a significant percentage of the variation(25%) in the measured abiotic data and were grouped together as principal component 2 inthe principal components analysis. Moreover, it was this factor (fertility-soil reduction) thataccounted for the greatest percentage of the variation in cellulose decomposition (57%) inthe multiple regression. This nutrient-soil reduction factor had high positive loadings forinterstitial phosphorus (as well as PO4), inorganic nitrogen (NH4 + NO3) and sulfide and ahigh negative loading for Eh (Table 2). Thus, this factor comprises both soil fertility andsoil reduction components and is positively associated with cellulose decomposition.
The effect of soil fertility on decomposition is well documented, although not necessar-ily consistent. Rybczyk et al. (1996) reviewed 24 publications that examined the effects ofnutrient amendments on rates of decomposition for wetland litter: eight showed enhancedrates of decomposition, seven showed variable effects depending on the environmental con-ditions, and nine elicited no effect. Generally, nitrogen rather than phosphorus limits litterdecomposition, although phosphorus can stimulate litter decomposition after primary nitro-gen limitations have been alleviated (see Rybczyk et al., 1996) and has been shown to limitleaf decomposition in some woodland stream systems (Elwood et al., 1981). Furthermore,phosphorus limits soil respiration (Amador and Jones, 1993) and the decomposition of cel-lulose (Maltby, 1985) in phosphorus deficient environments such as the Florida Everglades,where soluble P concentrations are exceptionally low (Koch-Rose et al., 1994). Althoughstudies investigating the effects of nutrient addition onPhragmitesdecomposition are few,decomposition rates were found to be 10% greater in an eutrophic lake compared to anoligotrophic one (Andersen, 1978). Although this result was for aboveground tissue, wewould expect the nutrient limitation to belowground decomposition to be similar.
The effect of oxygen availability on organic matter decomposition in wetland environ-ments, as for nutrient effects, is not universally consistent. Surface litter in wetlands gen-
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erally decomposes faster when flooded than unflooded or infrequently flooded (Brinson,1977; Davis and van der Valk, 1978; Bruquetas de Zozaya and Neiff, 1991; van der Valk etal., 1991; Neckles and Neill, 1994). This positive effect of flooding on surface litter decom-position has been ascribed to the maintenance of adequate soil moisture for microbial/fungalcolonization and activity. Decomposition of organic matter in anoxic flooded soils, however,often produces a different response. Neckles and Neill (1994), for example, found signifi-cantly lower (by 20%) belowground decomposition in a freshwater wetland when the soilwas flooded and strongly reduced (soil Eh:−100 to 0 mV) than non-flooded and relativelymore oxidized (soil Eh: 200 to 400 mV). Also, Pozo and Colino (1992), supporting thenegative effects of anoxia on decomposition, identified slower decomposition belowgroundcompared to aboveground for leaf tissue in a salt marsh in Spain. Within a wetland soilprofile, decomposition rates often decrease with increasing depth, as found in the presentstudy, probably as a result of more strongly reduced soils with depth (Maltby, 1988; Schip-per and Reddy, 1995). Controlled laboratory studies have supported this conclusion thatdecomposition is reduced in more oxygen deficient conditions (Gale and Gilmour, 1988).These results are consistent with the general assumption that organic matter decomposi-tion is slower in anaerobic environments (Webster and Benfield, 1986). However, even incontrolled laboratory investigations, differences between aerated and anaerobic rates of de-composition are often small. For example, (Godshalk and Wetzel, 1978), who documentedonly small differences in oxic and anoxic decomposition rates, proposed that nutrients andtemperature are much more important than anoxia in controlling aquatic plant decompo-sition. Webster and Benfield (1986), reviewing decomposition in aquatic environments,also noted that the effects of anaerobiosis on decomposition rate are not always inhibitory.Maltby (1988), for example, found greater cellulose decomposition in more flooded zonesof a hydroperiod gradient in a bottomland hardwood site associated with the lower Mis-sissippi River floodplain. He speculated that cellulose breakdown may not be inhibited byanaerobic conditions in the field because of adapted bacteria. Nichols and Keeney (1973)speculated that the faster breakdown of water milfoil under low dissolved oxygen condi-tions was because anaerobic bacteria require less nitrogen and are therefore not nitrogenlimited. This conclusion is supported by the data of Ferdele and Vestal (1980) who investi-gated mineralization of lignocellulose in sediments of an oligotrophic lake. After 50 daysof incubation under anoxic and aerobic conditions at 13.5◦C, these researchers found only6 to 13% lower decomposition in the anoxic treatment. However, with P addition, degra-dation in oxic versus anoxic treatments were not different. They concluded that nutrientavailability was the most important factor controlling cellulose mineralization in this arcticlake. These studies demonstrate that microbial activity and resulting decomposition ratesare not always impaired by anaerobic conditions. Even the generally lower decompositionrates with depth can be partially attributed to factors other than anoxia, such as differencesin soil fertility with depth (Schipper and Reddy, 1995). For example, the lower rates ofdecomposition below ca. 40 cm depth in the present study may have been due to a changein soil texture, and hence fertility, from silt/clay to primarily sand (personal observation,Hans-Henrik Schierup). Furthermore, when significantly lower decomposition rates occurunder flooding or anaerobiosis, the difference is not necessarily of a large magnitude. Theability of the roots of wetland plants to oxidize reduced soil (McKee et al., 1988; Havens,1997) may be one of the interacting factors that allow for relatively high rates of below-
396 I.A. Mendelssohn et al. / Aquatic Botany 64 (1999) 381–398
ground decomposition even when the soil is saturated or flooded. The soil oxidizing powerof Phragmitesroots can result in greater bacterial counts than in more reduced parts of thesoil profile as well as greater glucose metabolism (Andersen and Hansen, 1982).
In the present investigation, the multivariate analyses did not separate out the effects ofsoil fertility from the effects of soil reduction. Sites exhibiting the highest rates of below-ground cellulose decomposition had both more reduced soils (higher sulfide and lower Eh)and higher fertility (greater N and P) than sites with lower decomposition rates. Thus, wecan ask: Which of these variables are more important in controlling belowground cellulosedecomposition along the studied salinity gradient? Based on the literature cited above, itis logical to conclude that more biochemically reduced, high sulfide soil environments,although not necessarily being inhibitory to decomposition, would not likely accelerate it.The soil Eh at the marsh sites exhibiting the highest rates of decomposition ranged from−200 to−100 mV. Eh levels this low have been associated with lower belowground de-composition of organic matter in a fresh marsh in Canada (Neckles and Neill, 1994). Incontrast, it is reasonable to assume based on the many studies that have demonstrated anincrease in decomposition with greater soil and/or water fertility that nutrient differencesamong sites were important in controlling differential cellulose decomposition. Both nitro-gen and phosphorus had high and significant correlations with decomposition rate (r = 0.65,P< 0.01;r = 0.76,P< 0.01, respectively). In this case, we hypothesize that the stimulatoryeffect of higher fertility on decomposition counterbalanced any potential negative effects ofthe more reduced soil condition. In fact, the highly reduced soils may have been instrumen-tal in promoting a more fertile soil by inhibiting nitrification and loss of nitrogen througheither nitrate leaching or denitrification in the more reducing zones of the soil as well as byincreasing phosphorus solubility.
In summary, we suggest that the high spatial variation in cellulose decomposition amongthesePhragmites-dominated marsh sites was likely driven by differences in soil fertility,primarily nitrogen and phosphorus. Soil reduction was introduced into the multivariate anal-yses only because it was correlated (negatively) with inorganic nitrogen and phosphorusconcentrations in the soil interstitial water. Regardless, it is apparent from this study thatunder field conditions anaerobic conditions do not necessarily inhibit decomposition wheninorganic nutrient availability is high. Whether the decrease in cellulose decompositionwith depth is primarily a result of increased soil reduction or decreased nutrient availabilityrequires further investigation. Interstitial conductivity and the major cations (Na, Mg, andK) contributing to it did not appear to have a negative effect on cellulose decomposition.However, our interpretation of the data suggests a negative effect of seawater on cellulosedecomposition controlled by other mechanisms. Additional research is needed to addressthis point. Although the absolute rate ofPhragmitesdecomposition will differ from that ofcotton strip cellulose, based on our results we suspect that the soil organic matter decom-position dynamics ofPhragmiteswetlands will be substantially affected by environmentalproblems that increase eutrophication. In contrast, the effects of sea level rise on organicmatter decomposition are less clear, and although warranting considerable further investiga-tion, increases in salinity appear to have the potential for decreasing cellulose decompositionrates.
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Acknowledgements
Funding for this research was provided by the Department of Plant Ecology, AarhusUniversity, Denmark and the Environment and Climate Programme of the European Com-mission, contract no. ENV4-CT95-0147 (EUREED).. We thank the technical staff of theDepartment of Plant Ecology, Aarhus University for their assistance in this research.
References
Amador, J.A., Jones, R.D., 1993. Nutrient limitations on microbial respiration in peat soils with different totalphosphorus content. Soil Biol. Biochem. 25, 793–801.
Andersen, F.Ø., 1978. Effects of nutrient level on the decomposition ofPhragmites communisTrin. Arch.Hydrobiol. 84, 42–54.
Andersen, F.Ø., Hansen, J.I., 1982. Nitrogen cycling and microbial decomposition in sediments withPhragmitesaustralis(Poaceae). Hydrobiol. Bull. 16, 11–19.
Armstrong, J., Afreen-zobayed, F., Armstrong, W., 1996a.Phragmitesdie-back: sulphide- and acetic acid-inducedbud and root death, lignifications, and blockages within aeration and vascular systems. New Phytol. 134, 601–614.
Armstrong, J., Armstrong, W., van der Putten, W.H., 1996b.Phragmitesdie-back: bud and root death, blockageswithin the aeration and vascular systems and the possible role of phytotoxins. New Phytol. 133, 399–414.
Brinson, M.M., 1977. Decomposition and nutrient exchange of litter in an alluvial swamp forest. Ecology 58,601–609.
Bruquetas de Zozaya, I.Y., Neiff, J.J., 1991. Decomposition and colonization by invertebrates ofTypha latifolialitter in Caco cattail swamp (Argentina). Aquat. Bot. 40, 185–193.
Carpenter, S.R., Adams, M.S., 1979. Effects of nutrients and temperature on decomposition ofMyriophyllumspicatumL. in a hard-water eutrophic lake. Limnol. Oceanogr. 24, 520–528.
Cizkova-Koncalova, H., Kvet, J., Thompson, K., 1992. Carbon starvation: a key to reed decline in eutrophic lakes.Aquat. Bot. 43, 105–113.
Colberg, P.J., 1988. Anaerobic microbial degradation of cellulose, lignin, oligolignols, and monoaromatic ligninderivatives. In: Zehneder, A.J.B. (Ed.), Biology of Anaerobic Microorganisms. Wiley, New York, pp. 333–372.
Costanza, R., d’Arge, R., de Groot, R., Farber, S., Grasso, M., Hannon, B., Limburg, K., Naeem, S., O’Neill,R.V., Raskin, R.J., Sutton, P., van den Belt, M., 1998. The value of the world’s ecosystem services and naturalcapital.. Ecol. Econ. 25, 3–15.
Davis, C.B., van der Valk, A.G., 1978. The decomposition of standing and fallen litter ofTypha glaucaandScirpusfluviatilis. Can. J. Bot. 56, 662–675.
Elwood, J.W., Newbold, J.D., Trimblle, A.F., Stark, R.W., 1981. The limiting role of phosphorus in a woodlandstream ecosystem: effects of a P enrichment of leaf decomposition and primary producers. Ecology 62, 146–158.
Ferdele, T.W., Vestal, J.R., 1980. Lignocellulose mineralization by arctic lake sediments in response to nutrientmanipulation. Appl. Environ. Microbiol. 40, 32–39.
French, D.D., 1988. Patterns of decomposition assessed by the use of litter bags and cotton strip assay on fertilizedand unfertilized heather moor in Scotland. In: Harrison, A.F., Latter, P.M., Walton, D.W.H. (Eds.), CottinStrip Assay: An Index of Decomposition in Soils. Institute of Terrestrial Ecology, Grange-Over-Sands, UK,pp. 100–108.
Gale, P.M., Gilmour, J.T., 1988. Net mineralization of carbon and nitrogen under aerobic and anaerobic conditions.Soil Sci. Soc. Am. J. 52, 1006–1010.
Godshalk, G.L., Wetzel, R.G., 1978. Decomposition of aquatic angiosperms II. Particulate components. Aquat.Bot. 5, 301–327.
Good, R.E., Good, N. F., Frasco, B. R., 1982. A review of primary production and decomposition dynamics of thebelowground marsh component, Estuarine Comparisons pp. 139–157.
Harrison, A.F., Latter, P.M., Walton, D.W.H. (Eds.), 1988. Cottin Strip Assay: An Index of Decomposition inSoils, ITE Symposium no. 24. Institute of Terrestrial Ecology, Grange-Over-Sands, UK, 176 pp.
398 I.A. Mendelssohn et al. / Aquatic Botany 64 (1999) 381–398
Havens, K.J., 1997. The effect of vegetation on soil redox within a seasonally flooded forested system. Wetlands17, 237–242.
JMP Statistics and Graphics Guide. SAS Institute Inc., Cary, NC, 580 pp.Koch, M.S., Mendelssohn, I.A., 1989. Sulphide as a soil phytotoxin: differential responses in two marsh species.
J. Ecol. 77, 565–578.Koch-Rose, M.S., Reddy, K.R., Chanton, J.P., 1994. Factors controlling seasonal nutrient profiles in a subtropical
peatland of the Florida Everglades. J. Environ. Qual. 23, 526–533.Latter, P.M., Harrison, A.F., 1988. Decomposition of cellulose in relation to soil properties and plant growth. In:
Harrison, A.F., Latter, P.M., Walton, D.W.H. (Eds.), Cottin Strip Assay: An Index of Decomposition in Soils.Institute of Terrestrial Ecology, Grange-Over-Sands, UK, pp. 68–71.
Latter, P.M., Howson, G., 1977. The use of cotton strips to indicate cellulose decomposition in the field.Pedobiologia 17, 145–155.
Maltby, E., 1985. Effects of nutrient loadings on decomposition profiles in the water column and submerged peatin the everglades, IPS Symposium. Tropical Peat Resources – Prospects and Potential, pp. 450–464.
Maltby, E., 1988. Use of cotton strip assay in wetland and upland environments – an international perspective. In:Harrison, A.F., Latter, P.M., Walton, D.W.H. (Eds.), Cottin Strip Assay: An Index of Decomposition in Soils.Institute of Terrestrial Ecology, Grange-Over-Sands, UK, pp. 140–154.
McKee, K.L., Mendelssohn, I.A., Hester, M.W., 1988. Reexamination of pore water sulfide concentrations andredox potentials near the aerial roots ofRhizophora mangleandAvicennia germinans. Amer. J. Bot. 75, 1352–1359.
Mitsch, W.J., Gosselink, J.G., 1993. Wetlands. Van Nostrand Reinhold, New York, 722 pp.Mitsch, W.J., Mitsch, R.H., Turner, R.E., 1994. Wetlands of the old and new world: ecology and management. In:
Mitsch, W.J. (Ed.), Global Wetlands Old World and New. Elsevier, Amsterdam, pp. 3–56.Neckles, H.A., Neill, C., 1994. Hydrologic control of litter decomposition in seasonally flooded prairie marshes.
Hydrobiol. 286, 155–165.Nichols, D.S., Keeney, D.R., 1973. Nitrogen and phosphorus release from decaying water milfoil. Hydrobiol. 42,
509–525.Nowicki, B.L., 1994. The effect of temperature, oxygen, salinity, and nitrogen enrichment on estuarine
denitrification rates measured with a modified nitrogen gas flux technique. Estuar. Coast. Shelf Sci. 38, 137–156.Ostendorp, W., 1989. Die-back of reeds in Europe – a critical review of the literature. Aquat. Bot. 35, 5–26.Ostendorp, W.C., Krauss, M., Krumscheid-Plankert, P., Moret, J.L., Rollier, M., Schanz, F., 1995. Lake shore
deterioration reed management and bank restoration in some Central European lakes. Ecol. Engn. 5, 51–75.Pozo, J., Colino, R., 1992. Decomposition processes ofSpartina maritimain a salt marsh of the Basque Country.
Hydrobiol. 231, 165–175.Reice, S.R., Herbst, G., 1982. The role of salinity in decomposition of leaves ofPhragmites australisin desert
streams. J. Arid Environ. 5, 361–368.Rybczyk, J.M., Garson, G., Day Jr., J.W., 1996. Nutrient enrichment and decomposition in wetland ecosystems:
models, analyses and effects. Current Topics in Wetland Biogeochem. 2, 52–72.Schipper, L.A., Reddy, K.R., 1995. In situ determination of detrital breakdown in wetland soil-floodwater profile.
Soil Sci. Soc. Am. J. 59, 565–568.Steinke, T.D., Charles, L.M., 1986. In vitro rates of decomposition of leaves of the mangroveBruguiera
gymnorrhizaas affected by temperature and salinity. S. Afr. J. Bot. 52, 39–42.Tanaka, Y., Tezuka, Y., 1982. Dynamics of detriuts-attached and free-living bacteria during decomposition of
Phragmites communispowder in seawater. Jap. J. Ecol. 32, 151–158.Titus, J.G., 1988. Greenhouse Effect Sea Level Rise and Coastal Wetlands. United States Environmental Protection
Agency, Washington, D.C. Report No. EPA-230-05-86-013.Updegraff, K., Pastor, J., Bridgham, S.D., Johnston, C.A., 1995. Environmental and substrate controls over carbon
and nitrogen mineralization in northern wetlands. Ecol. Applic. 5, 151–163.van der Valk, A.G., Rhymer, J.M., Murkin, H.R., 1991. Flooding and the decomposition of litter of four emergent
plant species in a prairie wetland. Wetlands 11, 1–16.Viles, H., Spencer, T., 1995. Coastal Problems Geomorphology, Ecology and Society at the Coast. Edward Arnold,
London, 350 pp.Webster, J.R., Benfield, E.F., 1986. Vascular plant breakdown in freshwater ecosystems. Ann. Rev. Ecol. System.
17, 567–594.Williams, M., 1993. Wetlands: A Threatened Landscape. Blackwell, Oxford, 419 pp.
